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SOUTHERN CALIFORNIA ACADEMY OF SCIENCES

Volume 114

TIN

Number 1

114(1) 1-62 (2015)

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April 2015

Southern California Academy of Sciences

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Date of this issue 15 July 2015

® This paper meets the requirements of ANSI/N!SO Z39.48-1992 (Permanence of Paper).

Bull. Southern California Acad. Sci.

114(1), 2015, pp. 1-11

Possible Stock Structure of Coastal Bottlenose Dolphins off Baja
California and California Revealed by
Photo-Identification Research

R.H. Defran,1’* Marthajane Caldwell,2 Eduardo Morteo,3,4 Aimee R. Lang,5 6 Megan

G. Rice,7 and David W. Weller5

1 Cetacean Behavior Laboratory, San Diego State University, 11060 Delphinus Way,

San Diego, CA 92126, USA

2 Marine Mammal Behavioral Ecology Studies Inc., 8429 Cresthill Avenue, Savannah,

GA 31406, USA

3Instituto de Ciencias Marinas y Pesquerias, Universidad Veracruzana, Calle Hidalgo
#617, Col. Rio Jamapa, C.P. 94290, Boca del Rio, Veracruz, MX
4Instituto de Investigaciones Biologicas, Universidad Veracruzana, Av. Dr. Luis
Castelazo Ayala SIN, Col. Industrial Animas, C.P. 91190, Xalapa, Veracruz, MX
5Marine Mammal & Turtle Division, Southwest Fisheries Science Center, National
Marine Fisheries Service, National Oceanic and Atmospheric Administration, 8901 La
Jolla Shores Drive, La Jolla, CA 92037, USA
6 Ocean Associates, Inc., 4007 North Abingdon Street, Arlington, VA 22207, USA
1 California State University, San Marcos, 333 S. Twin Oaks Valley Rd., San Marcos,

CA 92078, USA

Abstract. — Boat-based photo-identification research has been carried out on bottle-
nose dolphins in eastern North Pacific coastal waters off northern Baja California,
Mexico and southern and central California, USA from 1981 to 2001. Within these
waters, bottlenose dolphins routinely travel back and forth between coastal locations
while generally staying within a narrow corridor extending only 1-2 km from the
shore. Inter-area match rates for 616 dolphins photo-identified between 1981-2000 in
four California coastal study areas (CCSAs) of Ensenada, San Diego, Orange County
and Santa Barbara averaged 76%. To explore possible southern range limits for these
dolphins, photo-identification surveys were carried out in the coastal waters off San
Quintin, Baja California, Mexico between April- August 1990 (n= 8 surveys) and July
1999 to June 2000 («= 12 surveys). The 207 individual dolphins identified off San
Quintin were compared to the 616 dolphins identified in the CCSAs. The inter-area
match rate between San Quintin and the CCSAs was 3.4% (n=l dolphins). This low
rate contrasts sharply with the much higher average match rate of 76% observed
between the CCSAs. These differences in match rates suggest that both a California
coastal stock and coastal Northern Baja California stock may exist, with only
a limited degree of mixing between them.

The common bottlenose dolphin ( Tur slops truncatus) is the most frequently
encountered cetacean in the nearshore waters of California and Baja California, Mexico.
Two distinct bottlenose dolphin ecotypes occur in these waters: a coastal form that is
typically found within 1-2 km of shore (Carretta et al. 1998; Defran and Weller 1999;

* Corresponding author: rh.defran@gmail.com

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JUL 2 1 2015

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SOUTHERN CALIFORNIA ACADEMY OF SCIENCES

Bearzi 2005) and an offshore form that is distributed in deeper waters, typically greater
than a few kilometers from shore (Defran and Weller 1999; Bearzi et al. 2009).
Differentiation of these two ecotypes, which are managed as separate stocks by the
National Marine Fisheries Service (Carretta et al. 2013), is supported by morphological
(Walker 1981; Perrin et al. 2011), photographic (see Shane 1994) and genetic data
(Lowther-Thieleking et al. 2014).

The California coastal stock is small, estimated to contain about 450-500 individuals
(Dudzik et al. 2006; Carretta et al. 2013) that are distributed between Monterey,
California and Ensenada, Baja Mexico (Defran et al. 1999; Hwang et al. 2014), with
occasional sightings as far north as San Francisco, California1. Photo-identification
research has been carried out on the coastal stock off California, and to a lesser extent off
Northern Baja California, since the early 1980s. Areas off California and Baja California
where photographic data have been collected include: (1) Ensenada, (2) San Diego,
(3) Orange County, (4) Santa Monica Bay, (5) Santa Barbara, (6) Monterey Bay and
(7) San Francisco Bay (Fig. 1). In general, photo-identification data have shown that
California coastal dolphins display little site fidelity to any portion of their distribution
(Defran et al. 1999; Hwang et al. 2014). Instead, they routinely travel back-and-forth
within their range, on some occasions in excess of 900 km, while at the same time
typically staying very near shore (Defran et al. 1999; Hwang et al. 2014).

Records from the nineteenth century suggest that coastal bottlenose dolphins may have
once occurred in Monterey Bay and San Francisco Bay (Dali 1873; True 1889; Orr 1963).
More recent studies, however, considered the northern range boundary to be located off
Los Angeles County up until the early 1980s (Norris and Prescott 1961; Dohl et al. 1981;
Leatherwood and Reeves 1982). The 1982-83 El Nino Southern Oscillation (ENSO)
dramatically impacted the coastal marine ecosystem off California and Baja. It was during
this ENSO event that California coastal stock dolphins extended their northern range back
to Monterey Bay (Wells et al. 1990). This northern range extension has persisted to the
present day (Riggin and Maldini 2010; Maldini et al. 2010; Cotter et al. 2011) and now
extends even further north to San Francisco Bay and most recently, Bodega Bay1.

The southern boundary of the California coastal stock is less well known but photo-
identification data demonstrate that it extends to at least Ensenada (Defran et al. 1999;
Hwang et al. 2014). In this research, boat-based photo-identification surveys of coastal
bottlenose dolphins were carried out south of Ensenada off San Quintin Bay, Baja
California (Figs. 1 & 2). The goal of this research was twofold: (1) to examine the degree
of overlap between coastal dolphins photo-identified off San Quintin and those photo-
identified in study areas off Ensenada, San Diego, Orange County, and Santa Barbara,
and (2) to use photo-identification data to determine if the southern range of the
California coastal stock extended as far south as the San Quintin area.

Methods

The general design used in this study was the same as our earlier studies that compared
independently collected bottlenose dolphin photo-identification catalogs from California
and Baja California (Defran et al. 1999; Hwang 2014).

1 Szczepaniak, I., W. Keener, M. Webber, J. Stem, D. Maldini, M. Cotter, R.H. Defran, M. Rice, G.
Campbell, A. Debich, A. Lang, D. Kelly, A. Kesaris, M. Bearzi, K. Causey, and D. Weller. 2013.
Bottlenose dolphins return to San Francisco Bay. Poster presented at the 20th Biennial Conference on the
Biology of Marine Mammals, Dunedin, New Zealand December 9-13.

BOTTLENOSE DOLPHIN STOCK STRUCTURE, SAN QUINTIN, BAJA CALIFORNIA

Fig. 1 . Coastal locations where California coastal stock bottlenose dolphins have been photo-identified.
Point Conception and Punta Colonet are included to indicate the northern and southern coastal boundaries
of the Southern California Bight. Study areas marked with an asterisk indicate those that were compared to
San Quintin sightings (Table 1).

Study Area

This study was conducted in the coastal waters south of San Quintin Bay, Baja
California, during two independent study periods: 1) April, June and August 1990, n= 8
surveys (Caldwell 1992); and 2) July 1999 to June 2000, n= 12 surveys (Morteo et al.
2004). The San Quintin study area was located approximately 376 km south of San Diego
and about 200 km south of Ensenada. Within the study area, the survey track extended

SOUTHERN CALIFORNIA ACADEMY OF SCIENCES

32 km southward from a point 8 km east of Azufre Point (30o23’50” N; 115°54,42” W)
south to Rosario Canon (30°09’06” N; 115°48’27” W) (Fig. 2). Most surveys in 1990
began at the Base Camp near El Socorro and extended 18 km south to Rosario Canon.
During the 1999-2000 study period, surveys began about 8 km east of Azufre Point and
extended 26 km south to Hondo Creek (Fig. 2).

Photo-Identification Surveys and Photographic Data Analysis

Survey methodology and photo-identification analysis procedures employed during
both study periods in San Quintin followed those used previously in the Ensenada, San
Diego, Orange County and Santa Barbara study areas, hereafter referred to as California
coastal study areas (CCSAs). Detailed descriptions of these procedures are provided

BOTTLENOSE DOLPHIN STOCK STRUCTURE, SAN QUINTIN, BAJA CALIFORNIA

Table 1. Summary information on survey effort, study period, photographic data, and data sources
for all study areasa.

Study area

Number of surveys
(complete, partial)

Study period

Number of dolphins
identified

San Quintin

20 (20, 0)

19901, 1 999-20002

207

Ensenada

23 (23, 0)

1985-19863, 1999-20004

129

San Diego

241 (157, 84)

1981-19895, 1996-19996&7

518

Santa Barbara

73 (55, 18)

1987 & 19893, 1998-19997

213

Data sources: 1 Caldwell (1992), 2Morteo et al. (2004), 3Defran et al. (1999), 4 Guzon-Zatarain (2002),
5Defran and Weller (1999), 6Dudzik (1999), 7 Lang (2002). aSome numbers differ from those given in
original data sources due to refinement and revision of the dataset over time and the elimination of
sightings not meeting the specified photographic quality criteria.

elsewhere (Caldwell 1992; Defran and Weller 1999; Defran et al. 1999; Dudzik 1999;
Lang 2002; Morteo et al. 2004) but are briefly described here. Photographic surveys
involved slow travel in small boats while moving parallel to the coast and outside the surf
line; generally within 500-750 m of shore and corresponding to water depths between 4 m
to 10 m. Surveys were conducted in sea state and visibility conditions adequate for
finding and photographing dolphins. Although past data demonstrated that most coastal
bottlenose dolphins are typically found within 500 m of the shore (Hanson and Defran
1993; Defran and Weller 1999; Bearzi 2005; Carretta et al. 2013), two or more observers,
nevertheless, visually searched the area from the shore to ~ 2 km offshore to ensure
complete coverage of coastal waters. Once a group of dolphins was sighted, initial
estimates of group size, as well as information on time, location, environmental
conditions and behavior were recorded.

Following initial estimates of group size, the survey vessel maneuvered to a distance
from the dolphins suitable for photo-identification. Thirty-five millimeter SLR film
cameras equipped with telephoto lenses were used to photograph all dolphins (marked
and unmarked) within a group. Initial estimates of group size were revised as necessary,
and contact with the group was maintained until photographic effort was completed, or
dolphins began exhibiting avoidance behavior. Identical procedures were repeated as the
vessel resumed travel on the predetermined survey route and as additional dolphin groups
were encountered.

The best quality photograph of every dolphin was scanned and converted into a high-
resolution digital image. Of these, only high quality photographs of dorsal fins with two
or more distinctive dorsal fin notches were used for analysis. Distinctive dorsal fins were
those that had sufficient notching on the leading or trailing edge such that they could be
matched to high quality dorsal fin photographs from other sightings (Urian and Wells
1996; Defran and Weller 1999; Defran et al. 1999; Mazzoil et al. 2004). Only
unambiguous matches were accepted as resightings (i.e., a re-identification of a previously
identified individual). Dorsal fin images from selected CCSAs (marked with an asterisk in
Fig. 1) were analyzed and maintained in the Cetacean Behavior Laboratory at San Diego
State University. The combined photo-identification catalog for the CCSAs consisted of
616 individuals identified during two sample periods: (1) 1981 to 1989, and (2) 1996 to
2000. Table 1 provides a summary of survey effort, study period, photographic data and
data sources for each of the CCSAs.

SOUTHERN CALIFORNIA ACADEMY OF SCIENCES

120

110

100

1990

1999-2000

Study Period

Fig. 3. Number of dolphins identified during the 1990 and the 1999-2000 San Quintin study periods.
Some dolphins («= 9) were sighted during both study periods, otherwise individuals were sighted only
during the indicated study period.

Results

During the 1990 and 1999-2000 study periods in San Quintin, 104 and 112 individuals
were identified, respectively. Nine dolphins were identified in both the 1990 and 1999—
2000 study periods while 95 of these individuals were sighted only during 1990 and 103
only in 1999-2000. The combined number of individuals identified in San Quintin during
both study periods was 207 (Fig. 3). During the 1990 and 1999-2000 study periods, most
individuals were sighted only one time; but some individuals were sighted on multiple
occasions within a respective study period (Fig. 4).

Inter-study area match rates (MR) were derived by calculating the percent of
individuals photographed in one study area, such as in San Quintin, that were also
photographed in another study area. Similar match rate calculations were made for
individuals photographed within the different CCSAs. The first comparison involved
a composite of inter-study area matches reported for the two sampling periods when data
were collected in the CCSAs: 1981-1989 (Defran et al. 1999) and 1996-2000 (Hwang et al.
2014) (Table 1). Match rates for the 1981-1989 sample were calculated by comparing the
percent of dolphins identified in Ensenada (n= 68, MR =88%), Orange County (t?=133,
MR=92%) and Santa Barbara (n= 43, MR=88%) that matched to dolphins identified in
San Diego (n= 404) where the sample size was highest. Match rates for the 1996-2000
sample were calculated by comparing the percent of dolphins identified in Ensenada
(w=81, MR =43%) and Santa Barbara («= 182, MR =67%) that matched to San Diego
(n= 292) where the sample size was again the highest. The combined 1981-1989 and
1996-2000 average match rate for the CCSAs was 76% (±18.5 S.D.).

The second comparison involved the inter-study area match rates between dolphins
identified off San Quintin with dolphins in the combined 1981-1989 and 1996-2000
CCSAs catalog. Inter-study area matches occurred between both San Quintin datasets and

BOTTLENOSE DOLPHIN STOCK STRUCTURE, SAN QUINTIN, BAJA CALIFORNIA

Fig. 4. Sighting frequency of dolphins during the 1990 and 1999-2000 San Quintin study periods.

the CCS As catalog. In the 1990 San Quintin sample, 2 of the 104 dolphins identified were
matched (MR=1.9%) to the combined CCSAs catalog. In the 1999-2000 San Quintin
sample, 5 of the 112 dolphins identified were matched (MR=4.5%) to the combined
CCSAs catalog. When dolphins identified off San Quintin in 1990 and 1999-2000 were
combined (h=207), 7 (MR =3.4%) were matched to dolphins in the CCSAs catalog.

Finally, 3 of the 7 dolphins matched between San Quintin and the CCSAs catalog were
sighted in at least one of the CCSAs before and after their sighting(s) in San Quintin. The
first of these dolphins was sighted before and during the 1990 San Quintin study period
and again during the 1996-2000 CCSAs study period. The other two dolphins were
sighted after the 1996-2000 CCSAs study period, during more recent surveys conducted
in the San Diego study area between 2004—2011 (unpublished data, D. Weller). Of these
seven matches, all were sighted in San Diego, two were also sighted in Ensenada, and two
were also sighted in Orange County.

Discussion

Identifying population stock boundaries is important for management purposes in that
it allows for a range-wide evaluation of potential threats. With such management
considerations in mind, the most significant finding of this research was the low overlap
(MR =3.4%) for dolphins photographed off San Quintin and those photographed in one
or more of the CCSAs. In comparison, the overall match rate was considerably higher
(MR =76%) between CCSAs study sites. These match rate differences suggest that both

SOUTHERN CALIFORNIA ACADEMY OF SCIENCES

a California coastal stock and Northern Baja California coastal stock exist, with only
a limited degree of mixing between them.

Results from within the time frame of this research (i.e., 1990-1999) suggest that the
northern range boundary for the proposed coastal Northern Baja stock is located
somewhere between San Quintin and Ensenada. Although most individuals identified off
San Quintin were sighted only once, the small number of individuals that were sighted
multiple times within a given survey period provides some evidence for at least short-term
use of the area. Similarly, the nine dolphins sighted during both San Quintin study
periods indicate some degree of longer-term use of the area.

While the low match rate between San Quintin and the CCSAs suggests a small degree
of overlap between the two proposed stocks, the total number of surveys conducted off
San Quintin (n= 20) was relatively low in comparison to the number of surveys conducted
in some of the CCSAs. That being noted, the number of surveys conducted off San
Quintin (n= 20) is similar to the number of surveys off Ensenada (n= 23) that were also
conducted during two distinct time periods that overlapped or nearly overlapped with the
timing of the San Diego surveys. In this case, match rates for the Ensenada to San Diego
photo-identification comparisons were markedly higher (MR = 88% during 1981-1989;
MR =43% during 1996-2000) than those found for the comparison of the CCSAs catalog
with the two San Quintin survey periods (i.e., 1.9% and 4.5%, respectively). Further,
while the number of San Quintin surveys was lower than those in the San Diego and
Santa Barbara study areas, the number of dolphins identified was quite high. By way of
comparison, the 207 individuals identified (sampled) in San Quintin was greater than the
sample size in Ensenada (Table 1., n= 129) and comparable to the total number of
individuals in the Santa Barbara sample (Table 1., n= 213). Thus, it is unlikely that the
3.4% inter-area match rate between San Quintin and the CCSAs is related to low survey
effort or small sample sizes in San Quintin.

The primary variables contributing to the proposed stock structure are as yet
unknown. Oceanographic and bathymetric variables have been hypothesized as potential
habitat barriers for coastal bottlenose dolphins off California and Baja California
(Caldwell, 1992) but verification of these mechanisms is unresolved. When stock
separation occurs in bottlenose dolphins in the absence of confirmed geographic barriers,
as is the case along the eastern North Pacific coastline (this research), as well as along the
western North Atlantic Seaboard and within the northern inshore areas of the Gulf
of Mexico (e.g., Texas, Florida), social structure, prey availability, and foraging
specialization have been cited as possible foundations for dispersal tendencies (Sellas
et al. 2005; Rosel et al. 2009; Toth et al. 2011). Such stock distinctions may be useful for
management purposes, even when there is a moderate level of mixing with other adjacent
stocks, such as that which occurs within Sarasota Bay and between nearby Gulf of
Mexico inshore areas of Tampa Bay and Charlotte Harbor (see reviews in Selas et al.
2005; Rosel et al. 2009).

Complex social structure may act to minimize dispersal due to the investment required
to build and maintain social bonds (Rosel et al. 2009). Social affiliations among
California coastal dolphins, however, are highly dynamic (Weller 1991). Dispersal may
also be limited in areas that have consistently high prey densities, allowing a population
to be sustained long-term within a limited range. However, the regular travel of
California coastal dolphins within their range suggests a patchy distribution of prey
species requiring frequent relocation (Weller 1991; Hanson and Defran 1993; Defran
et al. 1999; Ogle 2005; Hwang et al. 2014).

BOTTLENOSE DOLPHIN STOCK STRUCTURE, SAN QUINTIN, BAJA CALIFORNIA

Among the inshore bottlenose dolphins found in some Atlantic Seaboard and Gulf of
Mexico areas, as well as within Shark Bay, Australia, a number of foraging and resource
specializations have developed. Over time, such specializations as strand feeding, sponge
feeding and confinement to shallow water bays and estuaries for shark avoidance, could
result in geographic range restrictions that give rise to stock separation (Silber and
Fertl 1995; Connor et al. 2000; Sellas et al. 2005; Mann et al. 2008; Rosel et al. 2009).
Similar mechanisms that might restrict the range of California coastal dolphins have not
been observed.

Examination of the sighting records for the seven dolphins identified in both San
Quintin and at least one of the CCSAs suggests that the mixing for some of these seven
dolphins may not represent permanent immigration of California coastal dolphins into
the putative coastal Northern Baja stock. Three of these seven dolphins were seen in at
least one of the CCSAs both before and after their sightings in San Quintin. Thus, these
dolphins appear to have visited San Quintin but subsequently returned to their putative
range within the CCSAs. It is unknown whether such visits entail exploratory movements
in search of prey and/or if they represent a mechanism by which some gene flow between
the two stocks could be occurring. A final point relates to the 3.4% match rate reported in
this research. The similarly low match rates observed for the two San Quintin sample
periods (i.e., 1990=1.9%, 1999-2000=4.5%) suggests that the degree of mixing does
fluctuate, at least to a small degree. Additional research conducted over years and
varying oceanic conditions could provide a more sensitive measurement of dolphin
mixing between the San Quintin and Southern California Bight study areas.

Thus far, the proposed stock separation presented herein relies entirely on photo-
identification data. The differentiation of California coastal ecotype bottlenose dolphins
from the offshore ecotype has successfully relied on the multiple data types and sources,
including analyses of morphology, microbiology and genetics, as well as photo-
identification (Walker 1981; Perrin et al. 2011; Bearzi et al. 2009; Lowther-Thieleking
et al. 2014). Among these multiple data sources, genetic analyses have been particularly
revealing in efforts to define and differentiate the coastal and offshore bottlenose dolphin
stocks within California waters (Lowther-Thieleking et al. 2014), as well as in numerous
other regions (Sellas et al. 2005; Rosel et al. 2009; Waring et al. 2012). Similar genetic
data are needed, but remain to be collected from San Quintin coastal dolphins. Once such
data are available, a genetic comparison to the California coastal stock can be made
(Lowther-Thieleking et al. 2014). Combining genetic comparisons with future photo-
identification data would provide a broader and more informed foundation from which
management decisions can be made with regard to coastal bottlenose dolphins off
California and Baja California.

Acknowledgements

The authors wish to thank the many dedicated students, interns and colleagues on both
sides of the border that assisted with this work. Brittany Hancock-Hanser and Jim
Carretta provided constructive reviews of an earlier draft of this paper.

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BOTTLENOSE DOLPHIN STOCK STRUCTURE, SAN QUINTIN, BAJA CALIFORNIA

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Colorado River Delta, Mexico. Aquat. Mamm., 21(3):183— 186.

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truncatus) stocks based on environmental, physical, and behavioral characteristics. Mar. Mamm.
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Bull. Southern California Acad. Sci.

114(1), 2015, pp. 12-21

Removal Efforts and Ecosystem Effects of Invasive Red Swamp
Crayfish ( Procambams clarkii) in Topanga Creek, California

Crystal Garcia,1,2 Elizabeth Montgomery,2 Jenna Krug,2 and Rosi Dagit2*

1 Water shed Stewards Program, 1455 Sandy Prairie Ct., Fortuna, CA 95540
2Resource Conservation District of the Santa Monica Mountains, 30000 Mulholland

Hwy, Agoura Hills, CA 91301

Abstract. — Red swamp crayfish ( Procambarus clarkii) were first recorded in Topanga
Creek in 2001 . When the onset of drought in Southern California resulted in low flows
and warming water temperatures from 201 1-2014, the population rapidly increased.
Within the Santa Monica Mountains, P. clarkii has been linked to diminishing
numbers of California newt ( Taricha torosa ), a species of special concern (Kats et al.
2013). To address these concerns, a student-based citizen science program was
conducted from November 2013 through April 2014 to remove crayfish from a 200 m
reach of Topanga Creek. The following data was collected and compared between the
removal reach and an upstream, adjacent 200 meter non-removal reach (control):
water quality (temperature, salinity, pH, conductivity, dissolved oxygen, turbidity),
nutrient levels (nitrate, nitrite, ammonia, orthophosphate), benthic macroinvertebrate
community metrics, crayfish demographics and catch-per unit effort (removal reach
only). The results indicate that red swamp crayfish presence or removals do not affect
water quality or nutrient levels in Topanga Creek. However, benthic macroinverte-
brate communities were significantly different between reaches; the presence of
crayfish correlated with lower BMI abundance and species richness, higher proportion
of tolerant taxa, and lower feeding group complexity.

Red swamp crayfish ( Procambarus clarkii ) have spread far across the globe, posing an
invasive threat to freshwater species abundance and community diversity (Ficetola et al.
2011). Mediterranean wetlands, such as those found along the southern coast of
California, have been shown to be preferred habitat for P. clarkii in periods of drought
with reduced flows and increased water temperatures (Geiger et al. 2005). This
crustacean grows rapidly, maturing within three months after hatching, and can
reproduce twice a year in warm conditions (Barnes 1974; Vodopich and Moore 1999).
Large healthy females typically produce 600 viable young furthering their ability to
spread quickly (Barnes 1974; Vodopich and Moore 1999). Procambarus clarkii are
omnivorous consumers of an array of plant and animal matter such as macrophytes,
detritus, amphibian eggs and larvae, aquatic invertebrates, and small fish, thus affecting
the riparian food web on a polytrophic scale (Momot et al. 1978; Momot 1995; Stenroth
and Nystrom 2003). The generalist and predatory feeding habits of P. clarkii have been
linked to observed declines in macrophyte abundance (Feminella and Resh 2006;
Rodriguez et al. 2005), amphibian species richness and recruitment (Gamradt and Kats
2002; Cruz et al. 2006; Ficetola et al. 2011), and macroinvertebrate diversity (Correia
and Anastacio 2008).

* corresponding author: rdagit@rcdsmm.org

PROCAMBARUS CLARKII IN TOPANGA: REMOVAL AND ECOSYSTEM EFFECTS

Red swamp crayfish were detected in southern California as early as 1924 (Holmes
1924), but not observed in Topanga Creek until 2001 ( RCDSMM unpublished data).
Topanga Creek is the third largest coastal watershed (47 km2) draining into the Santa
Monica Bay. Freshwater systems in this region are critical habitat that support a number
of sensitive and endangered native aquatic species. Procambarus clarkii were the first
introduced fauna to become established and spread throughout Topanga Creek, and
remains the most abundant non-native invasive in the watershed. The population of
P. clarkii in Topanga Creek was initially suppressed by active removal efforts and
significant winter rain events and sufficient flows to reduce crayfish abundance (Kats
et al. 2013). Below average rainfall and low flows in 2011-2014 have facilitated the
extensive establishment of P. clarkii throughout Topanga Creek.

The population growth of P. clarkii in Topanga Creek raised concerns about possible
implications for two native species, the California newt ( Taricha torosa ), a California
species of special concern, and federally endangered southern steelhead trout ( Oncor -
hynchus my kiss). Data collected from Topanga Creek during snorkel and other visual
surveys (2001-2014) documented the spread and increased abundance of P. clarkii , as well
as provided direct observations of crayfish attacking newts ( RCDSMM unpublished data
2014). The interactions of crayfish and O. mykiss are less clear; however, since 2011 an
increased incidence of crayfish found in the diet of large (>25.4 cm) O. mykiss has been
observed (Krug et al. 2012).

Benthic macroinvertebrates (BMI) are an important food source for both P. clarkii and
O. mykiss (Angradi and Griffith 1990, Nystrom and Graneli 1996). Competition for food
resources and disruption of BMI community functionality is a potential concern. The
complexity of functional feeding groups (e.g., gatherers, filterers, scrapers, predators) can
be a measure of the functional integrity of BMI communities and a reflection of its
capacity to cycle nutrients (Wallace and Webster 1996). Disturbance to the benthic
community, such as the introduction of non-native fauna, can alter BMI community
composition and cause unanticipated changes in freshwater ecosystems (Covich et al.
1999). Changes in BMI abundance, diversity, and feeding group complexity can indicate
such community disturbance.

In Topanga Creek, drought induced low flows in 201 1-2014 resulted in isolated refugia
pools and reduced numbers of O. mykiss redds and young of the year1. However, P. clarkii
were able to successfully reproduce and inhabit the shallow riffles and fragmented reaches
inaccessible to O. mykiss. In September 2013, the Resource Conservation District of the
Santa Monica Mountains (RCDSMM), in conjunction with the Watershed Stewards
Program (WSP), launched a citizen science program that 1 ) removed crayfish from several
refugia step-pool habitats within a 200 meter reach of Topanga Creek, 2) measured
crayfish demographics (sex/length), and 3) monitored water quality (dissolved oxygen,
pH, salinity, conductivity, turbidity, water temperature), nutrient levels (nitrate, nitrite,
ammonia, phosphate), and BMI community metrics.

Materials and Methods

Topanga Creek (34° 6T1”N 118° 36’ 18” W, gradient 1 to 6%) is the main drainage of
a small coastal watershed (approximately 47 km2) located within the Santa Monica

1 Krug, J., R. Dagit, Stillwater Sciences, and J.C. Garza. 2014. Lifecycle monitoring of Oncorhynchus
mykiss in Topanga Creek, California. Final Report Prepared for CA Department of Fish and Wildlife,

Contract No. P0950013. January 2014.

SOUTHERN CALIFORNIA ACADEMY OF SCIENCES

Mountains National Recreation Area in southern California. The study reach consisted
of 400 continuous meters in Topanga Creek, starting at 3500 m (upstream from the
ocean) and ending at 3900 m. The study area included a downstream 200 m crayfish
removal reach (RR), and an upper 200 m non-removal reach (NRR; Fig. 1). Both reaches
were relatively uniform in geomorphological features, including a similar distribution
of pools, step-pools, runs, and riffles, substrate type, and percent canopy cover. No
introduced barriers of any sort were incorporated into the study reaches; however,
natural low-flow boulder barriers separated the RR from the NRR.

A total of ten volunteer crayfish removal events took place between September 2013 and
April 2014. Water quality, nutrient, and BMI samples were collected in both 200 m reaches
during removal events between November 2013 and April 2014. Crayfish were removed
throughout RR with 7.6 cm hot dog pieces attached to hemp strings. The presence of
federally listed O. mykiss prevented setting traps of any kind. Crayfish were counted, sexed,
and measured (cm) from the tip of the rostrum to the end of the tail in midline. Removed
crayfish were donated to a local wildlife rescue or used for educational purposes.

Water samples were collected from three pools within each 200 m reach an hour prior
to removal. Each site was tested for air temperature (mercury thermometer), salinity
(ATC 300011 SPER SCIENTIFIC salt refractometer), pH (Oakton pHTestr 30),
conductivity (Oakton ECTestrl 1), dissolved oxygen (DO) and water temperature (YSI 55
DO meter). All probes were calibrated within a week prior to the collection date. Nutrient
and turbidity sampling was conducted once a month from November 2013 through April
2014 at 3500 m, 3550 m, and 3600 m in RR and at 3700 m, 3800 m, and 3850 m in NRR.
Samples were tested for nitrate-N (ppm), nitrite-N (ppm), ammonia-N (ppm),
orthophosphate (ppm) and turbidity (NTU) within eight hours of collection using
LaMotte SMART3 colorimeter and LaMotte 2020we turbidity meter.

BMI samples were collected according to CA Rapid Bioassessment protocol2 in
November 2013, December 2013, February 2014, and April 2014 at three comparable
sites in RR and NRR. Each sample was composed of nine kicks into a 1-ft. wide D-frame
net (three transects and three kicks per transect). Samples were preserved in 95% ethanol
and processed within a month from collection date. BMI were identified to genus, or
lowest possible taxonomic level using a 40x magnification dissecting microscope.
P. clarkii was recorded but not included as a benthic macroinvertebrate for analysis. For
quality assurance, 10 percent of samples were randomly selected and re-identified by
a second processor. First and second identifications were compared and scored for
accuracy, resulting in an estimated error of 1.6%.

Paired t-tests were applied to determine any significant difference between the two
reaches in crayfish demographics, water quality, nutrient levels, and biotic integrity
metrics of BMI communities. Regression analyses were performed to compare water
quality metrics to crayfish removal and to analyze the relationship between catch per unit
effort and water temperature. Simpson’s Index of Diversity (Simpson 1949) was
calculated for each BMI sample and analyzed by paired t-test to compare biodiversity.
Simpson’s was also applied to samples categorized by functional feeding groups
(gatherers/filterers, scrapers, predators, or other) to compare feeding group complexity.
Southern Coastal California Index of Biotic Integrity (SCC-IBI; Ode et al. 2005) metrics

2 Ode, P.R. 2003. CAMLnet: list of California macroinvertebrate taxa and standard taxonomic
effort. Aquatic Bioassessment Laboratory, Rancho Cordova. Retrieved September 10, 2014 from http://
www.safit.org/ste.html.

PROCAMBARUS CLARKII IN TOPANGA: REMOVAL AND ECOSYSTEM EFFECTS

3700-3900m

3500-3700m

NEVADA

San

Francisco'?

Jose

Topanga Creek
Watershed

itlFORNh

Los Angeles*

Topanga Watershed

Crayfish sampling site

Topanga Creek Sampling Locations

1 in = 1 miles

Source:

Imagery - ESRI

Topanga Watershed - CalWater
Sampling Locations - RCDSMM
Projection: NAD 1983 Albers

**Distances are linear meters from the ocean

Fig. 1. Map of Topanga Creek Watershed and the crayfish study reaches (3500-3700 Removal Reach
(RR); 3700-3900 Non-Removal Reach (NRR)).

SOUTHERN CALIFORNIA ACADEMY OF SCIENCES

flj

Tfc

250

200

150

100

SEP 13 OCT 13 NOV 13 DEC 13 JAN 13 FEB 13
FEMALE ■ MALE

Fig. 2. Total number of male and female crayfish removed each month Oct. 2013 to Feb. 2014.

(number of EPT, Coleoptera and predator taxa, and percent tolerant, intolerant, non-
insect, and collector-gatherers + collector-filterers) were applied and scored for all
BMI samples.

Results

Ten volunteer removal events between September 2013 and April 2014 (203.25 person-
hours) resulted in the removal of 345 P. clarkii; 166 females and 179 males (Fig. 2). The
average length of crayfish removed was 7.61(±0.348 SE) cm. There was no significant
difference between male and female average length or number removed. The first event (9/21/
2013) resulted in the most captures with more than four times as many crayfish removed than
any proceeding month. The catch per unit effort (CPUE) in the study period November 2013
to February 2014 ranged from 0.1 to 3.0 crayfish per person per hour, and increased
significantly with warmer water temperatures (R2= 0.67, F= 12.27, /?<0.05). An increase of
approximately 0.26 CPUE was calculated for every 1°C increase in temperature (Fig. 3). The
comparison of water quality and nutrient concentrations between the RR and NRR showed
no significant differences, except for salinity. Salinity showed a statistical difference between
reaches (paired two-tailed, /(3)=-4.65, p<0.02). The NRR had higher salinity throughout
the course of the study, although levels in both reaches ranged from 0-2 ppm.

The four BMI samples collected from the NRR in November 2013, December 2013,
February 2014, and April 2014 contained a total of 645 individuals and 38 taxa.
The samples collected from the RR contained a total of 3,642 individuals and 51 taxa.
A total of four phyla were represented including Arthropoda, Annelida, Mollusca, and
Nematoda. BMI abundance was significantly higher (paired two-tailed r(3)=3.59,
p <0.04) in the RR (Fig. 4). In both reaches, there was an increase in BMI abundance
from November through April. The NRR had significantly lower richness (paired one-
tailed f(3) = 2.74, p< 0.04). However, species diversity as measured by Simpson’s Index of

PROCAMBARUS CLARKII IN TOPANGA: REMOVAL AND ECOSYSTEM EFFECTS

3.5
3

2.5
2

•a

3
u
<u

Oh 1.5

rt 1

0.5

R2 = 0.6715

11/12/13

L A

•12/17/13

„ ' '

11/26/13 , . - ■'*

*2/4/ 14' ” 1/21/14

,l/.7/'rf’ *

• 12/3/13

• 1/28/14

10 11 12 13 14

Water temperature (degrees celcius)

Fig. 3. Relationship between CPUE (catch/person/hour) and Water Temperature (°C) in Topanga
Creek Nov. 2013 to Feb. 2014.

Diversity (Simpson 1949) was not significantly different between sites and ranged from
0.66 to 0.84 for all samples.

In the RR, the three most dominant taxa were Chironomidae (midge larvae, 24%),
freshwater snails (Viviparidae and Hydrobiidae, 22% relative abundance), and Hyalella
(freshwater Amphipod, 15%) (Fig. 5). The three most abundant taxa in the NRR were

1400

1200

NOV 13 DEC 13 FEB 14 APR 14

REMOVAL — NON-REMOVAL

Fig. 4. Benthic macroinvertebrate abundance in samples collected from removal (RR) and
non-removal (NRR) Nov. 2013 to Apr. 2014.

SOUTHERN CALIFORNIA ACADEMY OF SCIENCES

Chironomidae
Hyalellidae
Ostracoda
Vivipaddae
< Physa

Coenagrionidae
Tnchoptera
Ephemeroptera
Pro cam barns darkii

0.1

REMOVAL

0.2 0.3 0.4

Relative abundance

NON-REMOVAL
0.5 0.6

Fig. 5. Eight dominant taxa collected in each 200m study reach.

Chironomidae larva (33%), Ostracoda (seed shrimp, 22%), and Hyalella (18%). The two
reaches shared the same six most dominant species, including the above mentioned with
the addition of Coenagrionidae (narrow-winged damselfly nymphs) and Physa (physa
snails). These dominant taxa described above each have a tolerance value of 8, with the
exception of Chironomidae, which has an assigned tolerance value of 6 although there is
great variation among genera and species.

While total SCC-IBI scores showed no trend, two SCC-IBI metrics differed significantly
between sites: % tolerant taxa and % collector-gatherer + collector-filterer. The NRR had
greater % tolerant taxa (tolerance values 8-10) than RR (paired two-tail /(3)=— 5.24,
/?=<0.02). The NRR had a greater proportion of collector-gatherer and collector-filterer
organisms (paired two-tail t(3)= -3.70,/?<0.04) and fewer scraper organisms (paired two-tail

REMOVAL

** % Collector-gatherer and -filterer

■ %Scraper

■ % Predator

■ % Other

NON-REMOVAL

% Collector-gatherer and -filterer
* %Scraper

■ % Predator

■ % Other

Fig. 6. Average functional feeding group composition in removal and non-removal reach samples
Nov. 2013 to Apr. 2014.

PROCAMBARUS CLARKII IN TOPANGA: REMOVAL AND ECOSYSTEM EFFECTS

/(3)=4.05, /?<0.03). In applying Simpson’s Index to sample data categorized by functional
feeding groups, functional feeding group diversity was significantly higher in the RR (two-
tailed, r(3)=3.41,/><0.05) (Fig. 6). Additionally, P. clarkii were collected more often in NRR
BMI samples (3.1%, 20 individuals total) than in RR (<1%, 7 ind.).

Discussion

The invasive Procambarus clarkii has been shown to have severe effects on native aquatic
wildlife in southern California streams (Riley et al. 2000, Gamradt and Kats 2002, Rodriguez
et al. 2005, Cruz et al. 2006, Feminella and Resh 2006, Correia and Anastacio 2008, Ficetola
et al. 2011). In Topanga Creek, benthic macroinvertebrate abundance and species richness
were significantly higher in the 200m RR where crayfish were actively managed by hand-
removal than in an adjacent NRR. This result is consistent with previous reports that
correlate non-native crayfish presence to reduced BMI abundance in freshwater systems
(Charlebois and Lamberti 1996, Stewart et al. 1998). In the RR, BMI samples contained
between 23 and 51 distinct taxa and in the NRR, richness ranged from 6-38. This finding
corroborates previous studies that have found that P. clarkii invasions lead to loss of BMI
diversity (Rodriguez et al. 2005, Correia and Anastacio 2008). Functional feeding group
diversity was lower in the NRR, and % of tolerant organisms was higher.

Increased abundance of BMI in RR indicates higher productivity for a number of taxa.
Six distinct taxa had more than 100 individuals in one or more samples from the
RR including Viviparidae and Hydrobiidae, Chironomidae, Hyalellidae, Coenagrionidae,
Ostracoda, and Physa. Only two taxa had more than 100 individuals in any one NRR
sample: Chironomidae and Ostracoda. A major distinction between community was that
Viviparidae and Hydrobiidae were most abundant taxa in RR, but relatively rare in NRR
(3%). The relative rarity of freshwater snails (scrapers) in the NRR diminished feeding
group complexity. Procambarus clarkii predation on Viviparidae in this reach is one
possible driver of reduced abundance of the genus, although micro-habitat differences
within the 400 m study reach are another potential factor. Higher abundance, species
richness, feeding group complexity, and a smaller proportion of tolerant species indicate
that the BMI community in RR was in better ecological condition than in NRR. As crayfish
are generally the largest species within the BMI community, a comparison of BMI sample
proportional dry weight of taxa groups would further our understanding of P. clarkia
effects on trophic-level productivity by providing a quantitative measure of biomass.

The ecological implications of invasive P. clarkii in Topanga Creek could be severe if
they significantly disrupt benthic macroinvertebrate communities. BMI make up the
primary consumer trophic level and play an integral part in nutrient decomposition and
cycling through riparian systems. Changes at this level could impact higher trophic
organisms such as California newts (species of special concern) and southern California
steelhead trout (endangered). How the continuation of drought conditions within the
region will continue to affect the population and impact of P. clarkii is uncertain; reduced
flows and higher temperatures place stress upon aquatic natives, it renders riparian
habitat more preferential for crayfish.

Water quality and nutrient results between reaches were less notable. Salinity was the
only parameter to differ significantly, which may be influenced by a groundwater seep in
NRR at 3900 m3. Some studies have suggested P. clarkii may be a source of bioturbation

3GeoPentech. 2006. Hydrogeologic Study Lower Topanga Creek Watershed, Los Angeles County, CA.
Prepared for the RCD of the Santa Monica Mountains. Topanga, CA.

SOUTHERN CALIFORNIA ACADEMY OF SCIENCES

(Mueller 2007, Yamamoto 2010), however, results in this study showed no significant
difference in turbidity between the RR and the NRR.

The level of effort per crayfish removed increased over the course of the study at a rate
that correlated to decreasing water temperatures. While decreased activity is one possible
factor, diminished crayfish numbers due to removal efforts is another. Removal events
might be most efficient in warmer months; however a more extensive study including
more removal areas and a longer time period is needed to determine whether there is
a relationship between temperature and catch per unit effort, as well as to more
completely characterize the effects of crayfish on water quality and the benthic
macroinvertebrate community in Topanga Creek.

Acknowledgements

We would like to extend a special thanks to the following individuals and organizations
for their contributions to this project: K. Vander Veen (Calvary Christian School), C. Najah
(Topanga Youth Wildlife Project), Daniel Paz (Verbum Dei intern), RCDSMM Stream
Team, Watershed Stewards Program, California Conservation Corps, and AmeriCorps.
Funding was provided by LA County District 3, Supervisor Zev Yaroslavsky. This paper
also benefitted from review by three anonymous reviewers.

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PROCAMBARUS CLARKII IN TOPANGA: REMOVAL AND ECOSYSTEM EFFECTS

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Bull. Southern California Acad. Sci.

114(1), 2015, pp. 22-32

Soil Organic Carbon and Nitrogen Storage in Two Southern
California Salt Marshes: The Role of Pre-Restoration Vegetation

Jason K. Keller,* Tyler Anthony, Dustin Clark, Kristin Gabriel, Dewmini
Gamalath, Ryan Kabala, Julie King, Ladyssara Medina, and Monica Nguyen

Schmid College of Science and Technology, Chapman University, Orange, CA

Abstract. — Soil organic carbon and nitrogen storage represent important ecosystem
services provided by salt marshes. To test the importance of vegetation on soil
properties, we measured organic carbon, total nitrogen, and belowground biomass
in two southern California salt marshes. In both marshes, cores were collected from
areas which differed in dominant vegetation cover prior to the restoration of tidal
influence. There were no differences in organic carbon or total nitrogen density
between vegetation classes at either site; however, a relationship between
belowground biomass and soil organic carbon suggests that vegetation may
influence soil properties.

Salt marshes provide a number of important ecosystem services, including habitat for
fish and bird species, food web support for adjacent marine environments, nutrient
removal from the landscape, and carbon storage in long-lived soil pools (e.g., Zedler and
Kercher 2005). Despite their importance, these ecosystems have been lost at alarming
rates. Recent estimates suggest that on a global scale, 25% of salt marshes have been lost
since the 1800s with ongoing loss rates of an additional 1-2% per year (Mcleod et al.
2011). While comparable estimates of loss rates in southern California are limited, it is
likely that salt marsh loss in the region is considerably higher than the global average.
Grossinger et al.1 used US Coast Survey T-sheets from the late 1800s to estimate
a historical area of 7,711 ha of vegetated intertidal marsh along the South Coast of
California (from Point Conception to the Mexico border). Sutula et al.2 estimated that
approximately 1681 ha (4,153 acres) of intertidal estuarine wetlands remain in the same
region. While a direct comparison between these values should be viewed with caution
due to differences in methodologies, the apparent dramatic loss in wetland area highlights
the impact of historical anthropogenic activities on Southern California wetlands.
More recently, losses of salt marsh habitat in the Pacific region were negligible between
2004-2009 (Dahl and Stedman 2013), suggesting that rates of loss have slowed. Further,
ongoing conservation and restoration activities are aimed at maintaining the services
provided by the remaining wetlands in the region (Callaway and Zedler 2009).

* Corresponding Author: jkeller@chapman.edu

Grossinger, R.M., E.D. Stein, K.N. Cayce, R.A. Askevold, S. Dark and A.A. Whipple. 2011.
Historical wetlands of the southern California coast: an atlas of US Coast Survey T-sheets, 1851-1889. San
Francisco Estuary Institute Contribution #586 and Southern California Coastal Water Research Project
Technical Report #589, 55 pp.

2 Sutula, M„ J.N. Collins, A. Wiskind, C. Roberts, C. Solek, S. Pearce, R. Clark, A.E. Fetscher,
C. Grosso, K. O’Connor, A. Robinson, C. Clark, K. Rey, S. Mrrissette, A. Eicher, R. Pasquinelli, M. May
and K. Ritter. 2008. Status of Perennial Estuarine Wetlands in the State of Califonia. Southern California
Coastal Water Research Project, 48 pp.

SOIL ORGANIC CARBON AND NITROGEN IN RESTORED SALT MARSHES

A great deal of recent attention has focused on capitalizing on ecosystem services
provided by salt marshes as a potential means to support ongoing restoration and
conservation efforts. In particular, there is a growing interest in quantifying carbon
storage in salt marshes (Chmura et al. 2003; Mcleod et al. 2011; Pendleton et al. 2012).
Salt marshes, along with other vegetated coastal ecosystems including mangroves and sea
grass beds, are particularly effective at storing carbon in their soils because anaerobic
conditions generally limit decomposition of primary productivity in these ecosystems
(Megonigal et al. 2004; Tobias and Neubauer 2009) while a continuous supply of sulfate
limits production of the greenhouse gas methane (Poffenbarger et al. 2011). Further, salt
marshes continuously accrete new soils vertically to cope with sea level rise, which allows
for new layers of soil carbon to be accumulated through time (Kirwan and Megonigal
2013; Morris et al. 2002). This so-called “blue carbon” could conceptually be traded
in emerging carbon markets, although there are a number of ecological, political and
economic questions surrounding this possibility (Edwards et al. 2013; Pendleton et al.
2013; Sutton-Grier et al. 2014; Ullman et al. 2013). Concomitant with storing “blue
carbon”, salt marsh soils serve as an important sink for nitrogen, and this ecosystem
service may also be valuable in the context of restoration and conservation efforts
(Lau 2013).

We have previously measured soil organic carbon storage in two restored salt marshes
in Huntington Beach, California (Keller et al. 2012). This work showed that soil organic
carbon was generally higher in a marsh that had been restored for two years than in an
adjacent marsh that had been restored for 22 years. This suggests that the assumption
that restoration projects share a common starting point and predictably accumulate soil
carbon through time needs to be critically evaluated. In particular, we hypothesized that
initial site conditions, such as extant vegetation, may be as important as time following
restoration when determining soil carbon storage, and perhaps when determining other
belowground ecosystem properties.

Here, we further explore this possibility by measuring soil carbon and nitrogen storage,
as well as belowground biomass, in two additional southern California salt marshes. In
the first marsh, which had been restored for three years, we compared belowground
properties from areas which differed in vegetation coverage prior to restoration. In the
second marsh, which had not yet been restored, we compared areas dominated by
dramatically different pre-restoration vegetation communities.

Materials and Methods

Site Description

The Huntington Beach Wetlands used for this project are remnants of a larger marsh
that historically existed at the mouth of the Santa Ana River in northern Orange County,
California (33° 39’ N, 117° 59’ W). The majority of this marsh area was isolated from
tidal exchange by the mid- 1940s due to development and flood control measures, but
various wetland restoration efforts, including reconnection to tidal exchange, have been
taking place since the 1980s3. To explore the importance of pre-restoration vegetation on
belowground carbon and nitrogen dynamics, we collected samples in both the Magnolia
and Newland Marshes (Fig. 1).

3 Jones & Stokes Associates, I. 1997. Talbert Marsh restoration project five-year postrestoration
monitoring report. Final. December. (JSA 96-300.) Sacramento, CA. Prepared for Huntington Beach
Wetlands Conservancy, Huntington Beach, CA.

SOUTHERN CALIFORNIA ACADEMY OF SCIENCES

Efforts to restore the 16.6 ha Magnolia Marsh, including reestablishment of tidal
influence as well as the recreation of historical tidal channels, were completed in 2010
(Gordon Smith, Huntington Beach Wetlands Conservancy, personal communication).
We utilized Google Earth images from October of 2007 to identify locations which
differed in pre-restoration vegetation coverage. Specifically, we collected four soil cores
from areas with vegetation cover prior to restoration (“vegetated”) and two soil cores
from areas with limited vegetation cover (“un vegetated”; Fig. 1C.). While admittedly
qualitative, our designations of vegetation cover are in general agreement with vegetation
monitoring efforts at Magnolia Marsh, which show extensive coverage of senescent salt
marsh vegetation on the eastern side and limited vegetation on the western side of this
site4. Core locations were not selected based on specific vegetation communities, but at
the time of collection vegetation was generally similar to other southern California salt
marshes, and included: pickleweed ( Salicornia pacified), alkali seaheath ( Frankenia
salina), turtleweed ( Batis maritima), and saltgrass ( Distichlis spicata).

At the time of our sampling, tidal influence had not yet been restored to the 17.8 ha
Newland Marsh, located west of Magnolia Marsh in Huntington Beach. This site
currently has two visually distinct vegetation communities; a salt marsh community (“salt
marsh”) dominated by plants similar to those found in Magnolia Marsh and a brackish
community (“brackish”) dominated by cattail ( Typha sp.) We collected two soil cores
from each vegetation community in Newland Marsh (Fig. ID.).

Sample Collection and Analysis

Soil cores were collected in October-December 2013 following a modification of the
protocol described in Keller et al. (2012). Briefly, a 15.3 cm diameter stainless steel tube
equipped with a sharpened bottom edge was inserted to an average depth of 41 cm below
the soil surface (range 32-48 cm). Care was taken to minimize soil compaction. Upon
extraction of the core, soils were sliced into 2 cm depth increments using a serrated knife
and returned to the laboratory at Chapman University for processing. Each depth
increment was weighed and then passed through a 2 mm sieve within 1 week of collection
(when necessary, soils were stored at 4°C until sieving). Material >2mm was subse-
quently washed with distilled water over a 1-mm sieve and live roots and rhizomes were
collected and dried at 60°C to a constant mass. Four depths from a core collected in the
brackish community at Newland Marsh had highly organic soils, which did not pass
easily through the 2 mm sieve. Belowground biomass was removed by hand from these
depths and the remaining (unsieved) soil was processed as described below. Subsamples
of soil that passed through the 2 mm sieve were dried at 60°C to determine percent
moisture for each depth increment. Percent moisture values were used to calculate the
total dry mass of soil based on the total wet mass collected at each depth. Dried soils were
ground to a fine powder using an IKA All Basic Analytical Mill (IKA Works, Inc.,
Wilmington, NC, USA). Organic carbon and total nitrogen were measured using
a Costech elemental analyzer (Costech Analytical Technologies Inc., Valencia, CA,
USA). To remove inorganic carbon, soil samples were acidified with 50 pL of 1M HC1
and dried overnight at 37°C twice before carbon and nitrogen analysis (Craft et al. 1991).

4Whitcraft, C., B. Allen and C. Lowe. 2013. Huntington Beach Wetlands Restoration Project
Monitoring Program Methodology and Data Summary. Prepared for Huntington Beach Wetlands
Conservancy and National Oceanic and Atmospheric Administration - MSRP.

SOIL ORGANIC CARBON AND NITROGEN IN RESTORED SALT MARSHES

Fig. 1. Location of Huntington Beach Wetlands (A.) and of Magnolia Marsh (outlined in red) and
Newland Marsh (outlined in blue) (B.). Soil cores were collected from areas identified as vegetated (n=4)
or un vegetated (n=2) prior to restoration in Magnolia Marsh based on imagery from 2007 (image date:
10/22/2007) (C.). Soil cores were collected from areas dominated by brackish (n=2) or salt marsh (n=2)
vegetation in Newland Marsh (image date: 4/16/2013) (D.). Image source: Google Earth for all panels.

Organic matter content was determined as loss on ignition (LOI) following combustion at
400°C for at least 10 hours.

Statistical Analyses

Organic carbon and total nitrogen concentrations were multiplied by the total dry mass
of soil to calculate the mass of organic carbon and total nitrogen in each depth increment.
These values were subsequently summed over the 0-10 and 0^10 cm depth increments
and expressed as organic carbon or total nitrogen densities (g cm-3) based on the total
volume of these depth ranges (Keller et al. 2012). The 0-10 cm depth increment includes
the majority of roots found in these sites while the 0^40 cm depth increment includes the
entirety of the soil core. In cases where soils cores did not extend to a depth of 40 cm, the
average elemental and mass values from the 3 deepest depth increments were used for all
missing depths to 40 cm. In 2 cores from Newland Marsh, this approach was used for the
38-40 cm depth increment. In a single unvegetated core from Magnolia Marsh, averages
were used for the 32-40 cm depth increments. A similar approach was used to calculate
total belowground biomass (g) in the upper 10 and 40 cm of each soil core.

SOUTHERN CALIFORNIA ACADEMY OF SCIENCES

% Organic Carbon % Organic Carbon

0 5 10 15 0 10 20 30 40

Fig. 2. Depth profiles of soil organic carbon content (A., B.) and total nitrogen content (C., D.) in
cores collected from Magnolia and Newland Marshes. In Magnolia Marsh, cores were collected from
areas identified as vegetated (n=4) or un vegetated (n=2) prior to restoration. In Newland Marsh, cores
were collected from areas dominated by Brackish (n=2) or Salt Marsh (n=2) vegetation.

Independent t-tests were used to compare organic carbon densities, total nitrogen
densities and belowground biomass in the 0-10 and 0^10 cm depth increments between
vegetated and unvegetated cores in Magnolia Marsh and between brackish and salt
marsh cores in Newland Marsh. All data were normally distributed; however, data
frequently failed to meet assumptions of equal variance between groups based on the
Levene’s Test. In cases with unequal variances, we used the more conservative t-test
output that did not make assumptions about equal variance (IBM Corp 2012).
Differences were considered significant at /?<0.05 for all t-tests. Regressions were used to
explore relationships between LOI, organic carbon and total nitrogen content as well as
relationships between soil organic carbon density and belowground biomass. All analyses
were completed using Version 21 of the IBM SPSS statistical package (IBM Corp 2012).

Results

Organic carbon content was highest in surface soils and decreased with depth at both
Magnolia Marsh and Newland Marsh (Fig. 2A and B.). Vegetated cores at Magnolia
Marsh had higher average organic carbon concentrations than unvegetated cores in the
upper 10 cm, but these differences disappeared at deeper depths (Fig. 2A.). Average
carbon density to a depth of 10 cm in vegetated cores at Magnolia Marsh was nearly
double the carbon density in unvegetated cores; however, there were no significant
differences in carbon density between vegetated and unvegetated cores over either the

SOIL ORGANIC CARBON AND NITROGEN IN RESTORED SALT MARSHES

Table 1. Mean (± 1 SE) soil organic carbon density, total nitrogen density and belowground biomass
in soil cores collected from Magnolia and Newland Marshes. All values were summed to a depth of either
10 cm or 40 cm. There were no significant differences between vegetated and unvegetated samples in
Magnolia Marsh or brackish and salt marsh samples in Newland Marsh at either depth.

Magnolia Marsh

Newland Marsh

Vegetated

(n=4)

Unvegetated

(n=2)

Brackish

(n=2)

Salt Marsh
(n=2)

Organic Carbon Density (g cm-3)
0-10 cm 0.023 ± 0.0014

0-40 cm °-013 ± °-0002

0.012 ± 0.0045
0.013 ± 0.0024

0.026

0.014

± 0.0092
± 0.0018

0.022 ± 0.0044
0.016 ± 0.0016

Total Nitrogen Density (g cm-3)
0-10 cm 0.0019 ± 0.00013

0^10 cm 0.0012 ± 0.00030

0.0012 ± 0.00003
0.0012 ± 0.00020

0.0018

0.0012

± 0.00060
± 0.00005

0.0020 ± 0.00030
0.0014 ± 0.00015

Belowground Biomass (g)

0-10 cm 12.9 + 4.0

0-40 cm 17.2 ± 5.4

4.5 ± 3.4
6.0 ± 2.1

11.8

19.0

± 11.2
± 18.1

6.0 ± 3.0
8.3 ± 2.8

0-10 or (M-0 cm depths (Table 1). Cores from the brackish community at Newland
Marsh generally had higher average organic carbon content than cores from the salt
marsh community, although variability between cores was high, especially in the brackish
community (Fig. 2B). There were no significant differences in organic carbon densities
between brackish and salt marsh cores in Newland Marsh when calculated over the 0-10
or CMO cm depths (Table 1). Patterns of soil nitrogen through the depth profile mirrored
organic carbon concentrations at both Magnolia and Newland Marshes (Fig. 2C and D),
reflecting a strong relationship between organic carbon and nitrogen in the soils. There
were no significant differences in total nitrogen density between vegetated and
unvegetated cores in Magnolia Marsh or between brackish and salt marsh cores in
Newland Marsh at either the 0-10 or CMO cm depth increments (Table 1).

Similar to organic carbon and total nitrogen, belowground biomass was generally
higher in surface soils and decreased with depth (Fig. 3). Average total belowground
biomass in both the 0-10 and 0-40 cm depth increments was nearly 3-times higher in the
vegetated cores compared to the unvegetated cores in Magnolia Marsh; however, these
differences were not statistically significant at either depth increment (Table 1). In
Newland Marsh, average total belowground biomass in both the 0-10 and 0^10 cm depth
increments was approximately twice as high in brackish cores compared to salt marsh
cores, but these differences were not significant at either depth range (Table 1). Total
organic carbon density in the 0-10 cm depth increased with increased belowground
biomass in the same depth range (/?=0.03; r2=0.48; Fig. 4). There was no relationship
between organic carbon density and belowground biomass in the 0^1-0 cm depth
increment (/?= 0.63; Fig. 4). Across all sites, organic carbon content increased with
increasing concentrations of organic matter (measured as LOI; /><0.001; r2=0.96;
Fig. 5A.). Similarly, total nitrogen content was highest in samples with high organic
matter content (p<0.001; r2=0.96; Fig. 5B.).

Discussion and Conclusions

Tidal influence had been restored at Magnolia Marsh for 3 years prior to sampling for
this project and had yet to be restored at the nearby Newland Marsh. Despite different
restoration histories, the upper 40 cm of soil in both sites stored between 0.01 3-0.0 1 5 g cm-3

SOUTHERN CALIFORNIA ACADEMY OF SCIENCES

Belowground Biomass (g) Belowground Biomass (g)

02468 10 12 14 0 2 4 6 8

Fig. 3. Depth profiles of belowground biomass in Magnolia (A.) and Newland (B.) Marshes. Cores
were collected as described in Figure 2.

of organic carbon (Table 1). These values are lower than the global average soil organic
carbon density of 0.039 ± 0.003 g cm-3 provided by Chmura et al. (2003). Soil organic
carbon density measured in the current project was also lower than the values of
0.034 g cm-3 and 0.023 g cm-3 measured in the adjacent Brookhurst Marsh and Talbert
Marsh which had been restored for 2 and 22 years, respectively (Keller et al. 2012). Taken
together, these results verify our previous assertion that time since restoration does
not appear to be the primary control of soil organic carbon content in this salt marsh
landscape. This conclusion is in contrast to previous chronosequence studies which have
documented increased soil carbon through time following restoration (e.g., Cornell
et al. 2007; Craft et al. 2003).

However, Streever et al. (2000) suggested that inter-site differences in ecosystem
properties may be greater than differences that emerge through time following
restoration. We previously hypothesized that site-specific differences in pre-restoration
vegetation may play a particularly important role in determining soil carbon density
(or other soil conditions) at these sites (Keller et al. 2012). The current project provides
limited support for this hypothesis. While there were trends towards higher soil carbon
and nitrogen in the vegetated cores in Magnolia Marsh and the brackish cores in
Newland Marsh (Fig. 2), these differences were not significant at either site (Table 1).
It is worth noting that there was considerable spatial variability in soil properties even
within a plant community type within the same marsh (especially in the brackish
community in Newland Marsh). The reasons for this variability are unclear, but could
include differences in marsh elevation, vegetation community and/or decomposition
dynamics which are known to interact to influence carbon content and rates of soil
accretion in marsh ecosystems (Kirwan and Megongial 2013). Future work should
consider this variability when attempting to account for carbon storage within an entire
marsh ecosystem.

Across both sites, 48% of the variability in soil organic carbon density in the upper 10 cm
was explained by belowground biomass in the same depth interval (Fig. 4), suggesting that
vegetation community can perhaps influence soil properties. Root and rhizome dynamics
are rarely studied in wetland environments due to logistical constraints (e.g., Iversen
et al. 2012), but these belowground processes may be important for understanding soil
carbon and nitrogen dynamics. Decreasing belowground biomass with depth has been
observed previously (Saunders et al. 2006) and may be driven by both biotic factors

SOIL ORGANIC CARBON AND NITROGEN IN RESTORED SALT MARSHES

Belowground Biomass (g)

Fig. 4. Relationship between soil organic carbon density and belowground biomass in the upper
10 cm (closed symbols) and the upper 40 cm (open symbols) of salt marsh soil cores collected from both
Magnolia and Newland Marshes.

(i.e., competition between species) and abiotic factors (i.e., flooding and oxygen
availability or their interaction). Modeling approaches have explored the links between
root productivity and soil carbon content (e.g., Mudd et al. 2009), and Langley et al. (2009)
demonstrated that organic matter production in the form of fine roots in response to
elevated atmospheric C02 was the primary driver of increased rates of accretion in
a brackish marsh.

There was a strong relationship between soil organic carbon content and organic
matter content (LOI) across all samples analyzed in the current project (Fig. 5). This
relationship was similar to those reported by Craft et al. (1991) and Callaway et al. (2012)
using salt and brackish marsh soils from North Carolina and San Francisco, California,
respectively (Fig. 5), suggesting that this relationship is relatively robust across climate
and vegetation types. The quadratic form of this relationship results from an increased
fraction of organic carbon in organic matter in soils with higher organic matter contents.
For example, organic matter from the 0-2 cm depth increment contained 42 ± 3 (mean
± 1 SE) percent carbon compared to 22 ± 3 percent carbon in organic matter from the
8-10 cm depth increment. These values are all below the 58% of organic matter predicted
to be carbon based on the van Bemmelen factor (commonly used to convert organic
matter to organic carbon) and are generally below the more recent estimate of 50%
carbon suggested by Pribyl (2010). The deviations from these values are particularly
pronounced in deeper (older) soils which might suggest that carbon is being lost from
organic matter through time, perhaps through microbial respiration or through export of
dissolved carbon.

SOUTHERN CALIFORNIA ACADEMY OF SCIENCES

Fig. 5. Relationship between organic matter content (measured as loss on ignition, LOI) and organic
carbon content (A.) and total nitrogen content (B.) in salt marsh soils collected from both Magnolia and
Newland Marshes. Previously published relationships from Craft et al. (1991) and Callaway et al. (2012)
are included for comparison.

Craft et al. (1991) also reported a relationship between total soil nitrogen content and
organic matter content (LOI), suggesting that relatively simple measurements of LOI
might provide indirect information on soil carbon and nitrogen. We also observed
a strong relationship between soil nitrogen content and soil organic matter content
(Fig. 5); however, our soils had a higher percent of soil nitrogen for a given organic

SOIL ORGANIC CARBON AND NITROGEN IN RESTORED SALT MARSHES

matter content (i.e., lower C:N) than those analyzed by Craft et al. (1991). Thus, while the
relationship between organic carbon and organic matter appears to be robust across
climates and vegetation types, the relationship between nitrogen and soil organic matter
may be much more site-dependent and generalized relationships should be viewed with
caution.

A lack of a consistent accumulation of soil organic carbon along a chronosequence of
southern California salt marshes (from pre-restored to 22 years post-restoration) suggests
that site-specific factors may be as important as time since restoration in controlling the
“blue carbon” accumulation in these systems. Pre-restoration vegetation, as either the
presence or absence of vegetation in Magnolia Marsh or as different vegetation
communities in Newland Marsh, also did not play the key role in determining soil organic
carbon (or total nitrogen) content in these marshes. However, a strong relationship
between belowground biomass and soil organic carbon means that vegetation does likely
play some part in determining soil properties.

Acknowledgements

We thank the School of Earth and Environmental Sciences within the Schmid College
of Science and Technology at Chapman University for funding this project as the
laboratory component of the Fall 2013 Ecosystems Ecology course. Angelina Delgado,
Justin Drzymkowski, Kaitlin Fuller, Nicolas Lapointe, Jacob Lopez, Daniel Moore,
Cassandra Oregel, Steven Pham, Jesse Simons, and Ryan Ugale provided valuable field
and laboratory assistance. Amber Garcia and Jocelyn Paez from Orange High School
were instrumental in completing the carbon and nitrogen analyses of these soils.
The Board of the Huntington Beach Wetlands Conservancy under the leadership of
Dr. Gordon Smith graciously provided access to these sites.

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Academy of Sciences, USA, 106:6182-6186.

Lau, W.W.Y. 2013. Beyond carbon: conceptualizing payments for ecosystem services in blue forests on
carbon and other marine and coastal ecosystem services. Ocean & Coastal Management, 83:5-14.

Mcleod, E., G.L. Chumra, S. Bouillon, R. Salm, M. Bjork, C.M. Duarte, C.E. Lovelock, W.H.
Schlesinger, and B.R. Silliman. 2011. A blueprint for blue carbon: toward an improved
understanding of the role of vegetated coastal habitats in sequestering C02. Frontiers in Ecology
and the Environment, 9:552-560.

Megonigal, J.P., M.E. Hines, and P.T. Visscher. 2004. Anaerobic metabolism: linkages to trace gases and
aerobic processes. Pp. 3 1 7-^424 in Biogeochemistry (W. H. Schlesinger, ed, ) Elsevier-Pergamon.

Morris, J.T., P.V. Sundareshwar, C.T. Nietch, B. Kjerfve, and D.R. Cahoon. 2002. Responses of coastal
wetlands to rising sea level. Ecology, 83:2869-2877.

Mudd, S.M., S.M. Howell, and J.T. Morris. 2009. Impact of dynamic feedbacks between sedimentation,
sea-level rise, and biomass production on near-surface marsh stratigraphy and carbon
accumulation. Estuarine, Coastal and Shelf Science, 82:377-389.

Pendleton, L., D.C. Donato, B.C. Murray, S. Crooks, W.A. Jenkins, S. Sifleet, C. Craft, J.W. Fourqurean,
J.B. Kauffman, N. Marba, P. Megonigal, E. Pidgeon, D. Herr, D. Gordon, and A. Baldera. 2012.
Estimating global "blue carbon" emissions from conservation and degradation of vegetated coastal
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Pendleton, L.H., A.E. Sutton-Grier, D.R. Gordon, B.C. Murray, B.E. Victor, R.B. Griffis, J.A.V.
Lechuga, and C. Giri. 2013. Considering "coastal carbon" in existing U.S. federal statutes and
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Poffenbarger, H.J., B.A. Needelman, and J.P. Megonigal. 2011. Salinity influence on methane emissions
from tidal marshes. Wetlands, 31:831-842.

Pribyl, D.W. 2010. A critical review of the conventional SOC to SOM conversion factor. Geoderma, 156:
75-83.

Saunders, C.J., J.P. Megonigal, and J.F. Reynolds. 2006. Comparison of belowground biomass in C3- and
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305-322.

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Tobias, C. and S.C. Neubauer. 2009. Salt marsh biogeochemistry - an overview. Pp. 445^192 in Coastal
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Bull. Southern California Acad. Sci.

114(1), 2015, pp. 33—41

Identical Response of Caged Rock Crabs (Genera Metacarcinus
and Cancer) to Energized and Unenergized Undersea Power
Cables in Southern California, USA

Milton S. Love,* 1* Mary M. Nishimoto,1 Scott Clark,1 and Ann Scarborough Bull2

1 Marine Science Institute, University of California, Santa Barbara, CA 93106
2 Bureau of Offshore Energy Management, 770 Paseo Camarillo, Camarillo, CA 93010

Increasingly, energy generation facilities (i.e., wave and wind) are being sited in
offshore marine waters. The electricity generated from these facilities is transmitted to
shore through cables carrying alternating (AC) or direct (DC) current. If DC is used, it is
converted to AC for the North American grid at onshore stations. While these currents
produce both electric and magnetic fields, only the magnetic field, here called an
electromagnetic field (EMF), is emitted from the cable. Some marine vertebrates and
invertebrates can detect EMFs (summarized in Normandeau et al. 201 11). However,
while it is clear that organisms can detect EMFs, less well understood is how these
animals respond behaviorally to this stimulus, and concerns have been raised regarding
how these organisms might interact with energized subsea cables1. Among fishes, a few
field or quasi-field studies have produced what appear to be minor or equivocal
responses. For instance, in a study of three species of elasmobranchs held in offshore
mesocosms and subjected to EMF, there were some statistically significant differences in
behavior; however these differences were inconsistent among individuals within a species2.
In other studies, migrating European eels (Anguilla anguilla ) in the Baltic Sea slowed, but
did not halt, their swimming speed around an energized cable (Westerberg and Lagenfelt,
2008), and the movement of a number of fish species did not appear to be affected by an
energized cable off Denmark3.

Along the Pacific Coast of the United States, fishers have also raised this issue4; one of
the specific issues is how crabs (which form major fisheries along the Pacific Coast) might
respond to energized power cables. There have been few studies on the behavioral
changes that invertebrates might show in the presence of EMF although a small
laboratory study implied that Dungeness crabs ( Metacarcinus magister) were attracted to

* Corresponding author: love@lifesci.ucsb.edu

1 Normandeau, Exponent, and T. Tricas, and A. Gill. 2011. Effects of EMFs from undersea power
cables on elasmobranchs and other marine species. U.S. Dept. Int., Bur. Ocean Energy, Management,
Regulation, and Enforcement, Pacific OCS Region, Camarillo, CA. OCS Study BOEMRE 2011-09.

2 Gill, A.B., Y. Huang, I. Gloyne-Philips, J. Metcalfe, V. Quayle, J. Spencer, and V. Wearmouth. 2009.
COWRIE 2.0 Electromagnetic Fields (EMF) Phase 2. EMF-sensitive fish response to EM emissions from
sub-sea cables of the type used by the offshore renewable energy industry. COWRIE Ltd. COWRIE-
EMF-1-06.

3 DONG Energy and Vattenfall A/S. 2006. Review Report 2005 The Danish offshore wind farm
demonstration project: Horns Rev and Nysted offshore wind farms environmental impact assessment and
monitoring. The Danish Offshore Wind Farm Demonstration Projects.

4 Pacific Fisheries Management Council (PFMC). 2010. Letter from PFMC to Federal Energy
Regulatory Council, dated 19 June 2010. Titled COMMENT Reedsport OPT wave Park Project, FERC
No. 12713. Accessed 11 December 2013. http://www.pcouncil.org/wp-content/uploads/Cmt_Reedsport_
OPT_FERC.pdf

SOUTHERN CALIFORNIA ACADEMY OF SCIENCES

a zone of high EMF and that crabs in some zones with elevated EMF levels were
somewhat more active than control animals5. Needed are studies that address how
organisms respond to an in situ energized submarine power cable. The presence of
energized and unenergized AC submarine cables in close proximity to one another off the
coast of southern California allowed us to conduct such an experiment on crabs.

The experiments took place off Las Flores Canyon (34°28’N, 120 02’W), southern
California, USA. Here several energized and unenergized submarine power cables,
identical in construction, lie unburied on the seafloor and extend to offshore oil and gas
platforms (Fig. 1). We selected two cables for this study; one was energized and the other
unenergized. The two cables run parallel to each other, perpendicular to shore, and are
approximately 7 m apart. Note that in an ongoing study we have determined that the
EMF around the energized cable dissipates to background levels at a distance of about
one meter.

We used stiff plastic perforated boxes (88 cm x 57 cm x 23 cm) that were secured to the
sea floor with sand anchors at a bottom depth of 10 m. Each box was placed so that one
end was in contact with one of the two cables. In all, twelve boxes were installed, six
adjacent to the energized cable and six adjacent to the unenergized one. The boxes were
installed at intervals of 2.5 meters along each cable, half on the east side and half on the
west side and these alternated from one side to the other (Fig. 1). To reduce the chances
of crabs visually sensing the cable, plastic panels were attached to the end of each box
closest to the cable and identical panels were attached to the boxes on the end farthest
from the cable. To further reduce the chances that the crabs could sense a difference
between the cable end and the noncable end, we also removed the common brown
macroalgae Pterygophora californica that occurs on the cables but does not live on the
adjacent sea floor.

With the boxes in place along the energized and unenergized cables, divers stocked each
with one adult crab of either Metacarcinus anthonyi or Cancer productus, for an
experimental trial. Each crab, which was randomly selected from a stock of legal-sized
crabs provided by a commercial crab fisherman, was dropped through a hinged hatch,
which was centered in the middle of the cage. One hour after emplacement, divers recorded
the position of the crab within the box by visually dividing it into two halves, the portion
closest to the cable being designated “near-half’ and that furthest from the cable “far-half’
(Fig. 1). A second diver then opened the box to record EMF values (in microteslas - pT)
with a handheld EMF detector (EMF 1390 from General Tools & Instruments). Readings
were taken on the floor of each box at the edge closest to the cable and on the floor of that
box furthest from the cable. The boxes were then leaving the crab in the box. Divers
returned 24 hours later to observe where the crabs were positioned in the boxes and
recorded EMF values. The crabs were then removed from the boxes and new, previously
untested, crabs inserted for the next trial. Four sequential, 24-hour trials comprised an
experiment. A total of four experiments were conducted in 2013 (10-14 June, 9-13
September, 30 September^- October, and 7-1 1 October). Crabs were selected randomly for
each box. Gender was recorded for each crab with exception of the first experiment.

The primary question we addressed in this study is whether crabs responded differently
to the two types (energized and unenergized) of cables. The observations made 1 hour

5 Wilson, C.S. and D.L. Woodruff. 2011. A preliminary study on the effects of electromagnetic fields on
the burial behavior and location of the Dungeness crab, Cancer magister. Pacific Northwest National
Laboratory, Prepared for the U.S. Dept. Energy, Contract DE-AC05-76RL01830, PNNL-20729.

CRABS AND UNDERSEA POWER CABLES

Fig. 1. The location of the energized and unenergized cables used in the experiments and the
orientation of six of 12 boxes. The distance between the cables, about 7 m, is not drawn to scale.

SOUTHERN CALIFORNIA ACADEMY OF SCIENCES

and 24 hours after the crabs were set in the cages were evaluated separately. We used the
generalized linear model (GLM) approach to determine if crabs along the energized cable
were found nearer or farther from the cable compared to crabs along a non-energized
cable. A crab’s position, in the half of the box near or far from the cable, was the response
variable. Given the binomial distribution of the response variable, a logistic regression
model was used with a logit link function.

We used JMP software to fit each GLM to the data by Firth bias-adjusted maximum
likelihood estimation of the parameter vectors6. The most complete GLM model analyzed
included the effects of experiment (1-4), trial (1^1) nested within experiment, side of cable
that the cage was set (west, east), and type of cable (energized and unenergized) as well as
the intercept. A likelihood-ratio Chi-square test evaluated the hypothesis that all the
model parameters were zero. We also examined a sequence of simpler GLM models to
identify the best-fit model that might include as few as one predictor. Akaike’s
information criterion (AIC) was used to select between candidate models.

To determine if the genders responded differently to the energized and non-energized
cables, we first added gender as a predictor in the complete GLM model using data from
all but the first experiment when gender was not recorded. We used the same method
above to determine the best-fit model. We also parsed the data by gender to determine if
either male or female crabs, separately, responded differently to the two types of cables.
Again, we used the same GLM approach described above to determine if cable type alone
or with the other explanatory factors had a significant effect on a male or female crab’s
position in a box.

The EMF at the end of the boxes closest to the energized cable ranged from a mean of
46.2 pT to 80.0 pT during the experiments, and the readings on the far end of the boxes
never exceeded 0.9 pT (Table 1). Along the unenergized cable, EMF did not exceed
0.2 pT in the near half or far half of the boxes during the experiments. A total of 192 crabs
were used in this study; 24 crabs in each of four experiments on each cable (Table 2). The
positions of all 192 crabs were observed 1 hour after emplacement. A total of eight crabs
were recorded as lost 24 hours after emplacement during the four experiments; three crabs
in boxes along the unenergized cable and five crabs along the energized cable. Escapement
was not possible and loss of crabs was likely due to predation by octopuses.

The crabs responded no differently in the boxes along the unenergized and energized
cables. Both 1-hour and 24-hours after the crabs were set in the boxes, there were no
apparent differences in the proportion of crabs near the two types of cable regardless of the
side of cable where the boxes were set (Fig. 2). For a given observation period, experiment,
trial nested within experiment, side of cable that the cage was set, and type of cable had no
significant effect on the position of crabs in the boxes as evident from the GLM that was not
significantly different from the intercept model (1 hour: n=192, -log likelihood =5.676,
X2= 11.351, DF= 17, p=0.838, AIC=295.901. 24 hours: n=184, -log likelihood =7.946,
X2= 15.892, DF = 17, p=0.532, AIC=281 .037). None of the GLMs that incorporated fewer
explanatory factors could predict with statistical significance the variability in crab
responses in the boxes next to the cables one hour or 24 hours after deployment.

The proportion of crabs near the two types of cables 24 hours after deployment was
highly variable across experiments regardless of side of the cable the box was set (Fig. 2).

6 Schwarz, C.J. 2013. Sampling, regression, experimental design and analysis for environmental
scientists, biologists, and resource managers. http://people.stat.sfu.ca/cschwarz/Stat-650/Notes/
PDFbigbook-JMP/.

CRABS AND UNDERSEA POWER CABLES

Table 1. Level of electromagnetic field (microteslas - pT) in those parts of boxes closest to unenergized
and energized cables as read one hour and 24 hours after crabs were inserted. EMF readings at the farthest
end of the boxes were <0.1(iT at the unenergized cable and <0.9 fiT at the energized cable.

The lower n in experiments 1 and 4 were due to the flooding of the housing containing the EMF meter
after the first day of observations, which led to failure of the devices. However, note that the energized
cable used in this experiments has been in continuous use for many years and did not fail during the course
of these studies.

Experiment

Cable Type

1 hr

24 hr

Unenergized

0.0

Energized

46.2

11.4

Unenergized

0.0

0.1

0.0

Energized

57.0

7.4

55.5

8.7

Unenergized

0.0

0.1

0.0

Energized

54.2

9.3

56.1

0.0

Unenergized

0.1

0.0

0.1

Energized

80.0

19.7

51.0

10.1

Combining the observations from the four experiments, the proportion of crabs found
close to the two types of cable changed little between the observations made one hour and
24 hours after the crabs were set in the boxes (Fig. 3). One hour after emplacement, 53%
(51 of 96) of the crabs set along the unenergized cable and 55% (53 of 96) of the crabs
along the energized cable were observed in the near-half of the boxes (Fig. 3). The log-
likelihood test of the GLM showed no cable-type effect on crab response (n=192, -log
likelihood=0.042, X2=0.084, DF=1, p=0.772, AIC=270.876). The AIC for this single-
factor model indicates that it is no worse fit of the 1-hour data than the GLM of all
explanatory factors. In comparison, 24 hours after emplacement 56% (52 of 93) of the
crabs set along the unenergized cable and 51% (46 of 91 of the crabs set along the
energized cable were in the near-half of the boxes (Fig. 3). Although a slightly greater
proportion of crabs were nearer the unenergized cable than the energized cable,

Table 2. Number and gender (F = female, M = males, Unk = unknown) of crabs used in four
experiments. Gender of crabs in experiment 1 was not determined. Loss of crabs between one hour and
24 hours was likely due to predation by octopuses.

Unenergized Energized

Unk

Total

Unk

Total

Grand Total

Experiment 1
1 hr

24 hrs

Experiment 2
1 hr

24 hrs

Experiment 3
1 hr

24 hrs

Experiment 4
1 hr

24 hrs

SOUTHERN CALIFORNIA ACADEMY OF SCIENCES

Unerwgized cable, 1 hr

j§jj£ fwar-lmlf
■ br-hsT

^ F?

11111b

14-,

12-

10-

Unenergized cable, 24 hrs

ggiw«Mur
■ far-fta*

Eai£

Energized cable, 1 hr

14-i ®5 n©ar4wiir 14-,

Energized cable, 24 hrs

S8t near -ha r

W#f*

East

SicJeorcafeie

Fig. 2. The number of crabs positioned in the near-half and far-half of boxes on the west side and east
side of the energized and unenergized cables by experiment, one hour and 24 hours after deployment.

cable type in the single factor model had no effect on crab response (n=184, -log
likelihood =0.266, X2=0.5318, DF=1, p=0.466, AIC=259.897). As was the case for the
1-hour observations, the proportion of crabs found close to the two types of cable did not
differ 24 hours after the crabs were set in the boxes.

Some of the crabs were found in the opposite half of the box when reexamined
24 hours later. Along the energized cable, 23.1% (21 individuals) of 91 crabs moved to the
half of the box that was closer to the cable from the half farther, and 27.5%
(25 individuals) moved to the half of the box farther from the half nearer. Along the
non-energized cable, 21.5% (20 of 93 individuals) moved to half of the box that was closer
to the cable, and 18.3% (17) moved to the farther half of the box. Movement of crabs
within the boxes between the one-hour and 24-hour observations is unknown.

The addition of gender to the complete GLM faired no better using data from
experiments 2-4 when gender was recorded (1 hour: n=144, -log likelihood =6.632,
V= 13.265, DF= 14, p=0.506, AIC=221.950. 24 hours: n=137. -log likelihood =7.136,

CRABS AND UNDERSEA POWER CABLES

Crab position, genders combined, experiments combined,
after one hour and 24 hours

near-half

Cable type

Fig. 3. The number of crabs positioned in the
unenergized cables after one and 24 hours.

Cable type

•-half and far-half of boxes adjacent to energized and

X2= 14.272, DF=14, p=0.430, AIC=21 1.300). None of the GLMs that incorporated
fewer explanatory factors could predict with statistical significance the variability in
crab responses in the boxes next to the cables one hour or 24 hours after deployment.

Specifically, cable type had no effect on a crab’s position in the boxes regardless of
gender (Fig. 4). One hour after emplacement, 54% of the females next to the unenergized
cable (26 of 52 crabs) and 50% of the females next to the energized cable (28 of 56) were
found in the near half of boxes (n=108, -log likelihood =0.080, A2 =0.160, DF=1,
p=0.689, AIC= 155.643). Twenty-four hours later, a slightly higher proportion of crabs
were found next to both types of cables, 58% of the females (29 of 50 crabs) next to the
unenergized cable were found in the near-half of boxes, whereas 52% of the females set
along the energized cable (27 of 52 crabs) were in the near-half. Again, the females
responded no differently to the two cable types (n=102, -log likelihood =0.190,
A2=0.380, DF=1, p=0.538, AIC= 146.285). Males also responded no differently to the
two cable types. One hour after emplacement, 65% of the males next to the unenergized
cable (13 of 20 crabs) and 50% of the males next to the energized cable (8 of 16) were
found in the near half of the boxes (Fig. 4). Although it appears that a greater proportion
of males were found nearer the unenergized cable than energized cable, cable type in the
single factor GLM had no statistically significant effect on male crab response (n=36,
-log likelihood =0.410, X2=0.820, DF=1, p=0.365, AIC=54.8330). Twenty-four hours

SOUTHERN CALIFORNIA ACADEMY OF SCIENCES

Crab position, by gender, summed across experiments,
after one hour and 24 hours

b) 24 hours

30-

b) 24 hours

20-

Cabf® type

Fig. 4. The number of female and male crabs positioned in the near-half and far-half of boxes adjacent
to energized and unenergized cables, one hour and 24 hours after eployment.

later, 50% of the males next to the unenergized cable (10 of 20 crabs) and 53% of the
males next to the energized cable (8 of 15) were found in the near half of the boxes
(n=35, -log likelihood =0.019, Y=0.038, DF=1, p=0.846, AIC=55.228).

Pacific Coast crab fishers have voiced several concerns regarding crabs and their
potential responses to the EMF generated by submarine power cables. These concerns
generally relate to whether crabs are either attracted to, or repulsed by, EMF. If either of
these occurs, crab migrations might be compromised and, more specifically, crabs might
not walk over a cable to reach a baited trap. While this experiment does not address all of

CRABS AND UNDERSEA POWER CABLES

these concerns, it does imply that these two crab species may not respond either positively
or negatively to the levels of EMF generated by this specific energized cable.

Literature Cited

SAS Institute Inc. 2013. JMP Pro 11.0.0.

Westerberg, H. and I. Lagenfelt. 2008. Sub-sea power cables and the migration behaviour of the European
eel. Fish. Manage. Ecol., 15:369-375.

Bull. Southern California Acad. Sci.

114(1), 2015, pp. 42-53

Recent Decline of Lowland Populations of the Western Gray
Squirrel in the Los Angeles Area of Southern California

Daniel S. Cooper1 and Alan E. Muchlinski2

1 Cooper Ecological Monitoring, Inc., 255 Satinwood Ave., Oak Park, CA 91377
2Department of Biological Sciences, California State University, Los Angeles, 5151
State University Drive, Los Angeles, CA 90032

Abstract. — We provide an overview of the distribution of lowland and otherwise
isolated populations of the western gray squirrel ( Sciurus griseus) in the Los Angeles
area of southern California, an area that has experienced a recent and ongoing
invasion by the non-native eastern fox squirrel (, Sciurus niger ), an urban-adapted
species introduced a century ago. Away from its strongholds in the western Santa
Monica Mountains, San Gabriel Mountains, and Santa Ana Mountains, the western
gray squirrel is resident locally in both the Santa Susana and the Verdugo
Mountains, in Griffith Park, in low hills at the eastern periphery of the San Gabriel
Valley and in Claremont, and along the Santa Ana River canyon near Yorba Linda.
It also persists east of the Los Angeles area in residential areas of Redlands and
Yucaipa, which as of 2014 are still outside the range of the eastern fox squirrel. Here
we document several gray squirrel extirpation events within its lowland range, and
discuss factors influencing its persistence and its extirpation.

The western gray squirrel ( Sciurus griseus) is a large tree squirrel native to forests of
the western United States and extreme northwestern Mexico, with the subspecies
S. g. anthonyi common and widespread in oak- and pine-dominated areas of the hills and
mountains of southern California (Wilson and Reeder 2005). In the Los Angeles area,
a region we define as extending from eastern Ventura County east through Claremont
and south through the coastal plain into Orange County to the base of the San Joaquin
Hills, it also occurs in human-modified habitats, including large city parks and golf
courses, where scattered trees, particularly conifers, provide year-round food and shelter.
It is one of two tree squirrels in the Los Angeles area, the other being the eastern fox
squirrel {Sciurus niger), a non-native introduced in the early 1900s, and now abundant
throughout much of the Los Angeles area of southern California (Jameson and Peeters
1988, King et al. 2010).

As discussed by Linders and Stinson (2007) western gray squirrels are closely tied to
oak and evergreen woodland, and serve two main roles in maintaining native woodlands:
they harvest and bury acorns throughout the woodland, and disperse the seeds and fruit
of various oak woodland component tree and shrub species, such as California bay
{Umbellaria calif ornica). They also forage heavily on truffle-like mycorrhizal fungi found
in leaf litter and loose soil, which aid oaks in fixing nitrogen and retaining water through
dry months. During foraging, western gray squirrels deposit the spores of these fungi
through their droppings, thus spreading them throughout the oak woodland and
promoting the health of its trees. Because of this close association with oaks, the presence

Corresponding author: dan@cooperecological.com

WESTERN GRAY SQUIRREL IN LOS ANGELES AREA

of western gray squirrels may serve as an indicator of oak woodland health. By contrast,
the eastern fox squirrel is highly generalist in its food sources, requires a much smaller
home range (becoming super-abundant in urban settings), and occurs in a much wider
array of habitats than S. griseus away from the major mountain ranges in the region
(Gatza 2011, Ortiz 2014).

The history and origin of western gray squirrel populations on the floor of the Los
Angeles Basin are poorly understood. Today, most lowland populations of S. griseus are
strongly associated with planted pines and other conifers, which may now be crucial
habitat elements for the species. It was presumably naturally present at lower elevations
when oak woodland (mostly Quercus agrifolia ) once covered large areas of now-
urbanized places like the San Gabriel Valley, a pattern shared by numerous lower
montane plant and wildlife species that are able to persist locally at lower elevations in
suitable areas of canyons and woodlands (Cooper 2011). Later, as the region developed,
populations of S. griseus may have retreated to large urban parks and more wooded
residential areas, where it persisted through most of the 1900s, including those at the base
of the San Gabriel Mountains foothills from Pasadena east into Claremont (an area
referred to as the “mesa” by early naturalists, e.g., Grinnell 1898). It is also possible that
they colonized these areas later by moving down from the surrounding foothills, or that
both patterns occurred, with isolated lowland populations “winking” out periodically,
replenished by animals from surrounding highlands. Whatever the history, in the years
between the late 1990s and the mid-2000s, S. griseus became scarce or altogether absent
within many of these same neighborhoods. Clear instances of its extirpation and
subsequent replacement - directly or indirectly - by the non-native eastern fox squirrel
are now well documented (e.g., Muchlinski et al. 2009, Guthrie 2009, King et al. 2010).

Sciurus niger became established in the neighborhoods surrounding the eastern Santa
Monica Mountains in the western Los Angeles Basin during the decades following its
introduction in 1904, it only arrived in the San Gabriel Valley around 1990, the east San
Gabriel Valley around 1998, and the Claremont area and Orange County in the early
2000s (Guthrie 2009, King et al. 2010). In recent years, S. niger has also colonized much
of urbanized Santa Barbara County (P. Collins, pers. comm.) and portions of San Diego
County, the latter also following an early introduction (King et al. 2010). Now virtually
ubiquitous throughout the Los Angeles area from the San Fernando Valley east to San
Bernardino County and south through Orange County, S. niger appears to still be absent
at several urban-edge locations at the margins of the Los Angeles area, including parks
and neighborhoods in Redlands and Yucaipa, San Bernardino Co. (Ortiz 2014); canyons
in the lower San Gabriel Mountain foothills from the Sunland-Tujunga area east through
Claremont (Gatza 2011), and along the Santa Ana River at Gypsum Canyon, near Yorba
Linda, Orange Co. (AEM, unpubl. data).

Only a handful of local naturalists have noticed this turnover, and few published data
exist on the range of S. griseus in the Los Angeles area prior to the arrival of S. niger.
Today, only a few populations of S. griseus remain away from the larger mountains
[typically below around 457 m (1500’) a.s.l.], with only a handful, at the far eastern
periphery being free of S. niger. To ensure the ecological integrity of these remaining
populations of S. griseus - and of their habitat patches - particularly in areas where
S. niger has not yet invaded (or at least where it is not completely dominant), it is
important that remaining populations of S. griseus be identified and recognized by
conservation agencies and organizations. Since the late 1990s, we (DSC and AEM) have
been making notes on the occurrence of S. griseus in the Los Angeles area, as described

SOUTHERN CALIFORNIA ACADEMY OF SCIENCES

Distribution of western gray squirrel ( Sciurus griseus)

JpF] Core WGS populations

# Persisting

Irregular/Unknown
O Extirpated

Freeways/major highways

/s/

Counties

0 5 10 20 mi. k

1 I I I I I I I I ^

| — I — I — I — | — I — I — I — | In

0 10 20 40 km

Data Source Credits: Basemap - World Shaded
Relief Copyright: © 2014 ESRI, Western gray
squirrel distribution data by Daniel S. Cooper
and Alan Muchlinski, map design by Jennifer
Mongolo, November 2014.

Fig. 1. Map showing current (2012-2014) range of western gray squirrel in the Los Angeles area.

below. This paper synthesizes findings from each of these efforts and provides detail on
a dramatic and ongoing ecological replacement of a native species by a non-native one.

Materials and Methods

Little published information exists on the current or historical distribution of the
western gray squirrel, so we relied on a variety of sources, including online museum
databases for specimen records (www.vertnet.org, last search conducted 21 October
2014), and field notes and recollections of a network of environmental professionals and
colleagues in the Los Angeles area. DSC conducted surveys of birds and vegetation in the
Puente and Chino Hills on the east side of the Los Angeles Basin for two years in the late
1990s (1997-1998; see Cooper 2000), and kept field notes of all sightings of S. griseus
from this area. AEM collected data on observations of S. griseus through an online
survey form (http://instructionall.calstatela.edu/amuchli/squirrelform2.htm), which has
received over 9000 visits since January 23, 2007, through field studies by four graduate
students (Lewis 2009, Gatza 2011, Erkabaeva 2013, Ortiz 2014), and through his own
observations within and east of the San Gabriel Valley.

DSC initiated a volunteer-based tree squirrel survey of Griffith Park in summer 2010;
with ten observers each searching up to five of 40 similarly sized survey blocks in and

WESTERN GRAY SQUIRREL IN LOS ANGELES AREA

around the park. The following year, DSC and volunteers spent 25 days in the park
between 8 August and 21 November 2011, gathering observations on foraging, breeding,
aggression displays and other behavior, and from 5 August 2011 to 11 July 2012, DSC
conducted a region-wide search for any remaining S. griseus populations in the lowland
Los Angeles area away from known occupied habitat. During this period, DSC posted
short articles and requests for information on local listserves (e.g., Pasadena Audubon
Society; various neighborhood “Patch” websites). Also, DSC and colleagues made site
visits (30 min - 2.5 hrs in duration) on 21 dates to 32 different locations in the eastern
Santa Monica Mountains, the west San Gabriel Valley, and in the Verdugo Mountains
and San Rafael Hills north of Glendale following up on reports and checking all
accessible lowland areas with appropriate habitat. To supplement these surveys, DSC and
colleagues installed motion-activated cameras during the same time period at Descanso
Gardens in the San Rafael Hills west of Pasadena (two near the upper portion of the
property bordering open space) and in the Verdugo Mountains (nine within canyons in
three areas: La Tuna Canyon, Cedarbend Canyon, and Whiting Woods; ibid).

Results

Within its core range in mountains at the periphery of the Los Angeles area, Sciurus
griseus is a conspicuous resident in canyons and oak groves, and appears to have little
contact with S. niger except at the immediate urban-wildland interface zone (Gatza 2011).
Below around 457 m (1500’) a.s.l., numerous subpopulations of S. griseus persisted into
the 1990s in areas between these major mountain ranges, and in some cases, well onto the
floor of urbanized areas, as summarized below, and in Figure 1 and Table 1.

Eastern Santa Monica Mountains! Griffith Park

Griffith Park, at the far eastern end of the Santa Monica Mountains, appears to
support the only large remaining population of the species in this range east of Sepulveda
Pass/Interstate 405, with approximately 25-50 individuals largely confined to two main
drainages (Western Canyon and Vermont Canyon). During intensive searches of
potential habitat in 2011 and 2012, no observations of S. griseus were made between
Sepulveda Pass and Cahuenga Pass (U.S. 101), an area that includes significant open
space at Franklin Canyon Park and elsewhere. However, we remain hopeful that
S. griseus may persist here, as we were unable to obtain access into the large Stone
Canyon Reservoir open space (Los Angeles Dept, of Water and Power) just east of
Sepulveda Pass near Bel Air, which supports apparently suitable habitat.

Santa Susana Mountains! Simi Hills

Located on the northwestern edge of the San Fernando Valley, these generally arid
ranges are dominated by low-growing chaparral and coastal sage scrub, with a small
number of permanent streams and oak woodlands, best developed in the former range.
Ecologically, the Simi Hills are more similar to the Santa Monica Mountains immediately
to the south than the San Gabriel and Sierra Madre ranges to the north, while the Santa
Susana Mountains reach higher elevations and feature more montane elements such as
bigcone douglas-fir ( Pseudotsuga menziesii) and extensive savannah dominated by annual
grassland and valley oaks ( Quercus lobata ). Based on field notes of local naturalists,
S. griseus is absent from the Simi Hills, but persists in the Santa Susana Mountains at
Browns Canyon and Devils Canyon (S. Bernal, 2012, in litt.), and possibly at O’Melveny
Park (sight record on 20 April 2014, CSULA web survey). Its historical status in either

SOUTHERN CALIFORNIA ACADEMY OF SCIENCES

Table 1 . Distribution of “lowland” populations of western gray squirrel in Los Angeles area (i.e., away
from major mountain ranges/foothills), 2014. Status: P = Persisting population; E = Extirpated; EFS =
Eastern fox squirrel.

Location,

Date of last

Region

Subarea

as applicable Elevation Status EFS?

record (Source)

East Santa Monica Mountains

Beverly Hills

<1200’

Yes

1975 (specimen, LACM 60617)

Griffith Park

<1200’

Yes

2014

Santa Susana Mountains

1400’

Yes

20141’2

West San Gabriel Valley

Verdugo Mountains

1800’

Yes

2012 (DSC, unpubl. data)

San Rafael Hills

1300’

Yes

2014 (Erkabaeva 2013)

San Marino

Huntington

600’

Yes

2010 (CSULA database)

Library

Lacy Park

600’

Yes

1976 (specimen, LACM 90234)

Mission Canyon3

600’

Yes

2012 (CSULA database)4

Northeast Los Angeles (Forest Lawn Glendale)

600’

Yes

1997 (DSC, unpubl. data)

East San Gabriel Valley

San Jose Hills

Industry Hills

600’

Yes

2014 (AEM, unpubl. data)

Bonelli Park area5

1000’

Yes

2014 (AEM, unpubl. data)

Walnut Creek Park

800’

Yes

2012 (DSC, AEM, unpubl. data)

Galster Park

600’

Yes

1998 (DSC, unpubl. data)

Cal Poly Pomona

800’

Yes

2009 (AEM 2009)

Via Verde Country

800’

Yes

—2000 (AEM, unpubl. data)

Club

Western Puente Hills6

Whittier/Hacienda

800’

Yes

1998 (DSC, unpubl. data)

Heights

Powder Canyon

800’

Yes

2005 (R. Erickson,
unpubl. data)

Eastern Puente Hills/Chino Hills

Tonner Canyon

600-800’

Yes

2014 (R. Hamilton, L. Schmahl,
via email)

Chino Hills State

1200’

Yes

2014 (A. Ing, pers. comm.)

Park

Pomona Valley/Claremont

RSABG7

1200’

Yes

2014 (AEM, unpubl. data)

Pomona College

1200’

Yes

2012 (AEM, unpubl. data)

Arlington Dr.

1200’

Yes

2012 (CSULA database)

Redlands/Y ucaipa

North of I- 10

Univ. of Redlands/

1500’

2014 (Ortiz 2014)

Sylvan Park

3rd St., Yucaipa

2600’

2011 (CSULA database)

South of I- 10

Ford Park

1600’

2014 (Ortiz 2014)

Prospect Park

1600’

2014 (Ortiz 2014)

Rossmont Dr.

2000’

2009 (CSULA database)

Hilltop Dr.

2200’

2012 (CSULA database)

WESTERN GRAY SQUIRREL IN LOS ANGELES AREA

Table 1. Continued.

Region

Location,

Subarea as applicable

Elevation

Status

Date of last

EFS? record (Source)

Orange County

Anaheim Hills

Oak Canyon Nature

800’

Yes 2013 (CSULA database)

Center

Santa Ana River Canyon

Yorba Reg. Park

400’

Yes 2014 (B. Leatherman, via email)

Canyon RV Park

400’

No 2014 (AEM, unpubl. data)

1 Includes sight record from O’Melveny Canyon Park in 2014 (CSULA web survey).

2S. Bernal, unpubl. data.

3 Includes oak woodland patches along Kewen, Canon and Encino Dr. at San Marino/Pasadena border.

4 This population was seen continuously through 2010; the 2012 report was likely a dispersing individual
from elsewhere.

5 Includes Mountain Meadows Golf Course.

6 We use State Route 57 as the east/west dividing line.

7 Rancho Santa Ana Botanic Garden, Claremont.

range is not known (a single specimen exists from “Oat Mountain” from 1969, LACM
47332), nor is the size of the extant population in the Santa Susana Mountains. A recent
(2014) observation of a roadkill S. griseus on U.S. 101 at Las Virgenes Canyon Rd.
(C. DeMarco, via email) suggests that colonization north from the Santa Monica
Mountains might be a possibility without the freeway and associated development as
a barrier. Sciurus niger is common in the Simi Hills, particularly at the urban periphery
(DSC, pers. obs.).

Verdugo Mountains! San Rafael Hills

Populations of S. griseus in both the Verdugo Mountains and the adjacent San Rafael
Hills are isolated from the San Gabriel Mountains to the north by Interstate 210 and
by dense residential development along Foothill Blvd. Despite searching promising
areas such as La Tuna Canyon Rd., Crescenta Valley Park and the Whiting Woods
neighborhood on the north slope of the Verdugos, and Descanso Gardens and Scholl
Canyon in the San Rafaels, we could not locate any individuals during observational
surveys in 201 1-12. However, in approximately three months operating motion-activated
cameras in 2012, we detected single individual S. griseus at two sites, one in Cedarbend
Canyon and one near Whiting Woods, confirming that the species persists in the Verdugo
Mountains. In the San Rafael Hills, Erkabaeva (2013) observed four S. griseus in a group
on one occasion at Descanso Gardens in La Canada, as well as several lone individuals
here during 2012. Later, a motion-activated camera that had been placed at Descanso
Gardens since 2012 recorded a single S. griseus in July 2014, indicating the persistence of
at least a small population here. Sciurus niger is very common throughout both the
Verdugo Mountains and San Rafael Hills, including within seemingly pristine habitat far
from development (DSC, pers. obs.).

Northeastern Los Angeles

The hilly residential neighborhoods of Los Angeles just south of the San Rafael Hills
(including Eagle Rock and Highland Park) appear to have also lost at least one lowland
population of western gray squirrels. Several individuals were observed in planted pines
in the upper portions of Forest Lawn Glendale on the Eagle Rock border in 1997 (DSC,

SOUTHERN CALIFORNIA ACADEMY OF SCIENCES

pers. obs.), but a check in 2011 (this study) revealed only S. niger. Lack of records from
considerable time afield in this region and no sightings by naturalists based at the Debs
Park Audubon Center in Highland Park (J. Chapman, pers. comm.) suggest that no
current population of S. griseus persists in this area, which includes the lowermost
portion of the Arroyo Seco.

PasadenalWest San Gabriel Valley

Small numbers of western gray squirrel occur at the northern/urbanized edge of
Pasadena/Altadena (c. 457 m a.s.l.), where S. niger is now abundant. One was observed in
2011 (DSC) in the courtyard of an abandoned U.S. Forest Service facility adjacent
to Hahamongna Watershed Park near the Pasadena/La Canada border, and according to
a local resident, up to four individuals, including a probable family group (in January
2011), have been recorded here in recent years (L. Paul, via email). Sciurus griseus is
occasional in the more wooded residential neighborhoods along the northern tier of
Altadena at the base of the mountains (e.g., near Eaton Canyon and Kinneloa Canyon),
but has apparently abandoned locales slightly downslope in denser residential areas,
including a former retirement facility (“The Scripps Home” at 2212 N. El Molino Ave.)
that had its mature trees removed prior to a redevelopment effort in summer 2011
(an action which apparently drove out S. griseus , fide L. Paul). More significantly,
a population of S. griseus that once occurred in remnant oak-walnut woodland amid
residential estates along Mission Canyon at the border of San Marino and Pasadena
(including Lacy Park) persisted to around 2012, with the last records (each of a single
individual) being along Kewen Drive in San Marino on several dates in 2010 (J. Garrett,
via email), and again in 2012 (M. Nakamura, CSULA web survey form). We also last
received reports from the nearby Huntington Library around the same time (three
separate sightings; T. Allison, CSULA web survey forms in 2008 and 2010; S. Claytor,
photograph in 2008). We know of no remaining population of S. griseus here or along the
lower Arroyo Seco south of Hahamongna/Devil’s Gate Dam.

East San Gabriel Valley! San Jose Hills

As in the west San Gabriel Valley, western gray squirrels occur widely in canyons and
locally in residential areas in the foothills on the northern tier of the east San Gabriel
Valley (e.g., above Monrovia, San Dimas and La Verne, >305 m a.s.l.). South of here, the
low range of hills in the eastern San Gabriel Valley referred to as the San Jose Hills
apparently serves as an ecological connection between the San Gabriel Mountains and the
Puente-Chino Hills, which then connect to the much larger Santa Ana Mountains to the
south (see Cooper 2000). Here, the species persists only at Bonelli Park (San Dimas) and in
the “Industry Hills” near La Puente, and several extirpations have been very recent
(e.g., observed by DSC at Walnut Creek Park in Covina in 201 1 but not sine e;fide AEM).

Puente-Chino Hills

Western gray squirrels occurred in multiple canyons and open space areas from
Diamond Bar and Rowland Heights west into Whittier and La Habra Heights, and south
into Brea, and Chino Hills State Park during the late 1990s (DSC, unpubl. data).
A population in Turnbull Canyon in the Whittier Hills (far western Puente Hills) was
apparently extirpated in the late 1960s following a major fire that burned many mature
oaks (J. Schmidt, in litt.), indicating that even by then some loss had occurred. By the late
2000s they had been extirpated west of Harbor Blvd., with replacement by S. niger,

WESTERN GRAY SQUIRREL IN LOS ANGELES AREA

including along Powder Canyon in Rowland Heights/La Habra Heights, where S. griseus
was present in late 2005 (1, R. Erickson, unpubl. data) yet absent by 2007 (DSC, unpubl.
data; fide L. Longacre). The latter location is particularly notable, as the canyon is
directly contiguous to hundreds of acres of natural habitat, has been protected as part of
the Puente Hills Landfill Conservation Authority, and has seen little if any land use
change in the past 20 years. A devastating fire in 2008 that burned most of Chino Hills
State Park resulted in the immediate loss of most western gray squirrel populations there,
with only a very small number of individuals persisting in oak woodland in the remote
center of the park, north of San Juan Hill (A. Ing, pers. comm.).

Pomona Valley! Claremont

While still present at Rancho Santa Ana Botanic Gardens and along the San Gabriel
foothills through the northern portion of Claremont (e.g., San Dimas Canyon, Marshall
Canyon, fide AEM), western gray squirrel has been recently extirpated from several
areas, and replaced by S. niger, within the city of Claremont to the south, including the
Claremont Colleges area (Guthrie 2009). There are apparently no historical or recent
records of S. griseus from the eastern Pomona Valley nor along the lower Santa Ana
River Valley upstream of Prado Dam.

Redlands! Yucaipa (San Bernardino County)

Western gray squirrels are widespread and conspicuous residents of the San
Bernardino Mountains. However, lowland populations away from the lower foothills
persist (as of 2014) at University of Redlands, Sylvan Park, Ford Park, and Prospect Park
(Ortiz 2014). The species has also been reported in the “Sunset Hills” area of Redlands
just south of Interstate 10 and in an apparently small area of Yucaipa (including Third
St.), where they are found in mature pines in a residential area (CSULA web survey).
These populations do not appear to be in contact with S. niger as of 2014, and are much
higher in elevation than other lowland sites discussed. However, because they are
persisting away from the main mountain ranges in what is still obviously lowland
(non-montane or foothill) habitat, we have included them here.

South Orange County

In contrast to the report by Pequegnat (1951) that the western gray squirrel was not
found in the Santa Ana Mountains, the species is present in several oak-filled canyons in
the Santa Ana Mountains (e.g., Trabuco Canyon, CSULA web survey and J. Ortiz, via
email; Modjeska Canyon/Tucker Wildlife Sanctuary, CSULA web survey; Whiting
Ranch Wilderness Park, R. Hamilton, via email). Additional reports to the CSULA web
survey locate western gray squirrels at the suburban-wildlands interface west of Lake
Elsinore. Whether they are recent (post- 1950s) arrivals to this range is not known.

Away from the Santa Ana Mountains, two small populations are known from Oak
Canyon Nature Center in the Anaheim Hills, and along the “Santa Ana River canyon”
where the Chino Hills meet the northern Santa Ana Mountains (AEM, unpubl. data;
B. Leatherman, via email). We know of no records from the San Joaquin Hills, where
S. niger has been present in residential areas since around 2010 (R. Erickson, via email).
Like much of San Bernardino (and Riverside) County, S. niger has only recently (late
1990s) penetrated Orange Co., but it is now widespread and common into Irvine
(D. Willick, via email).

SOUTHERN CALIFORNIA ACADEMY OF SCIENCES

Other Areas

DSC (unpubl. data) observed a small number of what appeared to be western gray
squirrels in pines at the golf course at the Palos Verdes Country Club near Malaga Cove
on the Palos Verdes Peninsula in the 1990s; in this same area in roughly the same time
period, a local naturalist observed what appeared to be a single individual in the same
area (R. Melin, via email). A recent visit to this area (October 2014) revealed that it still
supported a dense forest of coast live oak, toyon ( Heteromeles arbutifolia ), many mature
planted conifers and eucalyptus, and a riparian strip running through the golf course
(DSC, pers. obs.). And, whereas the eucalyptus plantation here has apparently been
established for more than a century (Gales 1988), due to the extreme isolation of this area
from any other known S. griseus populations, its coastal location, and the possibility that
this population derived from deliberately introduced individuals (or pertains to the
eastern gray squirrel, Sciurus carolinensis), we consider a Palos Verdes population to be
"hypothetical” for now until more information is uncovered that would support its
inclusion in the current range of the species.

Discussion

Our investigation into the distribution of the western gray squirrel in the Los Angeles
area elucidates its status as essentially a foothill species that is now rare and declining
below around 457 m elevation, particularly in areas where it has come into contact with
the eastern fox squirrel. Away from its main strongholds in the western Santa Monica
Mountains, the San Gabriel Mountains, and the Santa Ana Mountains, small, isolated
populations persist only in the Santa Susana Mountains, Griffith Park, the Verdugo
Mountains and San Rafael Hills, the San Jose Hills, the Chino Hills, at Rancho Santa
Ana Botanic Gardens in Claremont, and in Redlands/ Yucaipa. Based on local
naturalists’ observations, several lowland populations appear to have declined in the
past five years, including that in Bonelli Park, the San Rafael Hills, Chino Hills State
Park, and along the Santa Ana River Canyon near Yorba Linda. Invariably, extirpations
have occurred concurrently with colonization by the ubiquitous S. niger.

It is probably unlikely that truly extirpated, isolated lowland populations in the area
will re-develop on their own. Areas of recent extirpation (or near-extirpation, where
S. griseus is no longer resident but may occur irregularly) are typically separated from the
nearest presumed source population by more than a kilometer, and generally by dense
residential or urban development. Multi-lane freeways now provide formidable barriers
between these areas of extirpation and source populations of S. griseus. Remarkably,
animals do persist in a handful of lowland areas with very limited habitat, including the
Industry Hills in La Puente, which suggests that certain small, isolated subpopulations
may act as “refugia”, perhaps from pathogens that periodically sweep through larger and
more intact populations. Of course, these same refugia are vulnerable to their own
extinction events, and so are almost certainly temporary.

Erkabaeva (2013) demonstrated that the length of projected coexistence of the two
squirrel species in a given habitat fragment depends upon both the size of the habitat
fragment and the structure of the habitat within the fragment, with length of coexistence
associated with a higher diversity of food bearing tree species and coniferous trees.
Sciurus griseus had a high probability of going extinct within a relatively short period of
time (10 to 40 years) in small to medium-sized habitat fragments. The presence of the
S. niger in the same habitat brought about extinction in a shorter period of time.

WESTERN GRAY SQUIRREL IN LOS ANGELES AREA

Competition with other squirrel species has been suggested as a potential cause of
S. griseus decline (or a contributor to its current patchy distribution) in the region, but the
mechanisms involved in this relationship need further study. Extirpation sites generally
support very high densities of S. niger , yet this species simply occurs at higher densities in
general. Sciurus niger is highly urban-adapted, and occurs at all the sites where S. griseus
has vanished, and we have not confirmed a site where S. griseus has been extirpated and
where S. niger is completely absent. Still, King (2004) found few interactions among
S. niger, S. griseus, and even California ground squirrel ( Spermophilus beecheyi) in her
study area where all three occur in San Dimas, California (eastern Los Angeles Co.), and
Ortiz (2014) also observed very few aggressive interactions between S. niger and S. griseus
in her local study areas. Regardless of the mechanism, the loss of S. griseus in these areas -
and region-wide - may be associated with a profound ecological change and degradation
of seemingly healthy oak woodland and other habitat, particularly in wildland areas
where replacement has occurred (e.g., the Puente-Chino Hills).

Larger wildland areas where S. griseus is persisting in the presence of S. niger are of
particular interest because these appear to offer the basic habitat needs of both species,
at least for some period of time, and possibly in different areas of the landscape.
The discovery of nests of S. griseus well into protected open space such as in the rugged
Cedarbend/Whiting Woods area of the Verdugo Mountains (DSC, unpubl. data) and at
San Dimas Canyon Park (King 2004) suggests a pattern of edge-avoidance, possibly
related to increased competition with the eastern fox squirrel at the urban edge. However,
this pattern breaks down at sites like Fern Dell in Griffith Park, where S. griseus occurs
a few feet from houses and dense urbanization (DSC, unpubl. data). Here, supplemental
feeding or food provisioning may simply be “propping up” the population of S. griseus
which has also been aided by the abundance of planted trees providing additional food
sources (fruits and nuts). Although we have made a few direct incidental observations of
supplemental feeding (e.g., unshelled peanuts dropped at Fern Dell in Griffith Park being
carried off by S. griseus), it probably occurs widely. Other vegetative characteristics that
allow S. griseus to persist here include some amount of closed-canopy woodland
(or woodland-like groves of trees) with an open understory rich in non-woody debris and
leaf litter; older, mast-producing trees for food; and at least a few very tall trees for nest
placement (Linders and Stinson 2007), characteristics that still apply to many parks in the
region.

More proximate factors in the decline of S. griseus relevant in our study area include
death from injury and disease. Mortality from roadkill has been shown to be a major
(if localized) factor in squirrel deaths in studies in Washington state (Linders and Stinson
2007), and S. griseus is frequently detected as roadkill in the Los Angeles area (pers. obs.).
Many sites at the urban-wildland interface, including sites with documented S. griseus
extirpations have roads along a canyon bottom, making squirrels that live in low densities
and that forage on the ground particularly vulnerable. Other important causes of death
and/or population decline include necrotic mange (found in many populations of S. griseus
but oddly, apparently undocumented in the introduced S. niger in California, per King
2004); habitat quality decline from removal or disruption of the forest canopy due to
development, tree-cutting, or fire; soil trampling and compaction (which reduces the
biomass of fungi and perhaps other foods); and extreme natural events such as prolonged
drought, which work synergistically to wipe out small populations. However, considering
how modified the current habitat of many lowland S. griseus populations is (e.g., planted
pines on golf courses), habitat transformation would seem to be a relatively minor threat.

SOUTHERN CALIFORNIA ACADEMY OF SCIENCES

Based on the continuing trend of extirpation in the region, we consider all existing
lowland populations of S. griseus to be highly imperiled throughout the Los Angeles area.
We estimate one of the largest intact populations within the urban core of the region, that
at Griffith Park, at well under 50 individuals, and even here it is geographically limited
within the park itself, with most of the population in two adjacent canyons (DSC,
unpubl. data). Smaller, more isolated populations such as that at Rancho Santa Ana
Botanic Gardens and at various patches in the San Jose Hills are now “landlocked” by
freeways and urbanization and are probably much more imperiled; populations here and
the Chino Hills are now surrounded and infiltrated by S. niger (fide A. Ing), and they may
not be able to resist continued invasion by this species. In the case of Redlands/Yucaipa,
it is likely only a matter of time before S. niger colonizes and saturates the residential
areas and parks where S. griseus currently occurs alone.

Should re-introduction of S. griseus to lowland areas be attempted, we recommend this
be limited to large, protected areas of natural habitat; however, reintroduction into areas
where S. niger has already saturated the surrounding landscape and S. griseus has
disappeared, such as at Franklin Canyon Park in Beverly Hills or along the lower Arroyo
Seco in Pasadena, seems unlikely to succeed in the long term. Another possibility might
be the modification of large closed landfills that have trees with a significant amount of
closed canopy and that produce appropriate food items. We refer readers to Gatza (2011)
for information on a Habitat Suitability Model that would support S. griseus while not
being conducive to S. niger. Landfills within large urban areas often cover hundreds of
hectares, and modification of portions of these landfills with corridors between suitable
habitat fragments could provide new habitat for “lowland” western gray squirrels. We
would not recommend introducing individuals from outside into areas of continued
occurrence, such as Griffith Park, which would have the potential to introduce an
unknown pathogen into vulnerable, isolated populations.

Literature Cited

Cooper, D.S. 2000. Breeding birds of a highly-threatened open space: the Puente-Chino Hills, California.
Western Birds, 31:213-234.

— . 2011. Rare plants of Griffith Park, Los Angeles, California. Fremontia, 38(4): 1 8—24.

Erkabaeva, K. 2013. Habitat structure and extinction risk modeling of Sciurus griseus in long-term
coexistence habitats of southern California. M.S. thesis, California State Univ., Los Angeles.
Gales, D. 1988. Handbook of Wildflowers, Weeds, Wildlife, and Weather of the South Bay and Palos
Verdes Peninsula, Third Edition. FoldaRoll Company, Palos Verdes Peninsula, California.
Garrett, K. and J. Dunn. 1981. Birds of Southern California: Status and Distribution. Los Angeles
Audubon Society, Los Angeles.

Gatza, B.P. 2011. The effects of habitat structure on western gray squirrels and invasive eastern fox
squirrels. M.S. thesis, California State Univ., Los Angeles.

Grinnell, J. 1898. Birds of the Pacific Slope of Los Angeles County. Pasadena Academy of Sciences
Publication No. 11.

Guthrie, D. 2009. Suburban Squirrels. Chaparral Naturalist, 49(1), September/October 2009.

Jameson, E.W. and H.J. Peeters. 1988. California Mammals. Univ. of California Press, Berkeley, CA.
King, J.L. 2004. The current distribution of the introduced fox squirrel ( Sciurus niger ) in the greater Los
Angeles metropolitan area and its behavioral interaction with the native western gray squirrel
(Sciurus griseus). M.S. thesis, California State Univ., Los Angeles.

— , M.C. Sue, and A.E. Muchlinski. 2010. Distribution of the eastern fox squirrel ( Sciurus niger) in
southern California. The Southwestern Naturalist, 55(1):42M9.

Lewis, S.A. 2009. Factors that allow the native western gray squirrel ( Sciurus griseus) and the introduced
eastern fox squirrel ( Sciurus niger) to coexist in certain habitats within California. M.S. thesis,
California State Univ., Los Angeles.

WESTERN GRAY SQUIRREL IN LOS ANGELES AREA

Linders, M.J. and D.W. Stinson. 2007. Washington State Recovery Plan for the Western Gray Squirrel.

Washington Dept, of Fish and Wildlife, Olympia, 128+viii pp.

Muchlinski, A.E., G.R. Stewart, J. L King, and S.A. Lewis. 2009. Documentation of replacement of native
western gray squirrels by introduced eastern fox squirrels. Bull. So. Calif. Acad. Sci., 108:160-162.
Ortiz, J.L. 2014. Behaviors of the native western gray squirrel ( Sciurus griseus ) and the invasive eastern fox
squirrel ( Sciurus niger ) in Los Angeles and surrounding counties. M.S. thesis, California State
Univ., Los Angeles.

Pequegnat, W.E. 1951. The biota of the Santa Ana Mountains. Journal of Entomology and Zoology. Nos.
3 and 4.

Wilson, D.E. and D.M. Reeder, Editors. 2005. Mammal species of the world: a taxonomic and geographic
reference. Third Edition. Smithsonian Institution Press, Washington, D.C.

Bull. Southern California Acad. Sci.

114(1), 2015, pp. 54-57

A Young-of-the-Year Giant Sea Bass, Stereolepis gigas Buries
Itself in Sandy Bottom: A Possible Predator
Avoidance Mechanism

Michael C. Couffer1 and Stephanie A. Benseman2

lGrey Owl Biological Consulting
2California State University, Northridge

The adult giant sea bass, Stereolepis gigas, (GSB) is the largest teleost inhabiting
California’s shallow rocky reefs, attaining a length of about 2.3 m (7 ft) and a maximum
weight of around 256 kg (563 lbs) (Baldwin and Keiser 2008). They range from Humboldt
Bay, California to Oaxaca, Mexico, including the Gulf of California (Miller and Lea
1972). Adults consume a wide variety of prey and occupy rocky bottom habitat ranging
from approximately 7^10 m (25-130 ft) of water (Miller and Lea 1972) and can forage
over sandy bottom, away from rocky reefs (Baldwin and Keiser 2008). After their peak
commercial catch in 1932, at just over 1 14,000 kg, the population quickly crashed and
their numbers have remained depressed ever since; this has inhibited detailed research
(Pondella and Allen 2008).

Young-of-the-year (YOY) GSB pass through various color phases and morphological
changes during early development. These transitions help it to appear cryptic, while
hiding to avoid predators during a vulnerable stage of life. When less than 2.5 cm (1 in),
these YOY appear black with several small white spots around its face (Fig. 1). This
black stage has very large black dorsal and pelvic fins, with transparent pectoral, anal,
and caudal fins. The black juveniles morph through a “brown” stage, to a bright orange
fish (Fig. 2). The black dorsal fin changes to orange, while the enormous pelvic fins
remain black. Color expands outward to include half of the pectoral and anal fins, and
the entire tail remains clear. The white spots remain from the earlier stages, and small
black spots also appear (Pers. obs., and Benseman, unpublished data). These YOY
appear to frequent open, sand and mud-bottomed habitat between 2-30 m (7-100 ft) for
the first few months after settlement (Benseman, unpublished data).

During a focused SCUBA survey for YOY GSBs at Veteran’s Park in Redondo Beach,
Los Angeles County, California, Michael Couffer located a roughly 2.5 cm (1 in) long
orange juvenile GSB in 5.5 m (16.5 ft) of water, floating upright in the bottom of a shallow
sandy depression with its dorsal and pelvic fins closed. The bottom was clean sand without
surface detritus. When approached, the GSB raised its dorsal and pelvic fins and left the
depression, moving slowly within 30 cm of the bottom. The fish was photographed to
record the sighting time in image metadata, and followed from about a meter away to
acclimate the fish to human presence so that it could be photographed in profile.

After the fish had moved about 9 m, Mr. Couffer approached to half a meter to
photograph its spot pattern. The fish startled, and darted toward the bottom at an angle.
As the fish reached the bottom, it turned on its side at the last instant and buried into the
soft sand by undulating its body like a flatfish. It pushed its head beneath the sand and
undulated until the entire fish was buried in under three seconds. I took several photos of

Corresponding author: mikecouffer@gmail.com

A POSSIBLE PREDATOR AVOIDANCE MECHANISM

Fig. 1. An 18mm YOY Giant Sea Bass from Newport Beach, California.

Fig. 2. A 75mm YOY Giant Sea Bass from Newport Beach, California.

SOUTHERN CALIFORNIA ACADEMY OF SCIENCES

Fig. 3. Scales on the side of the buried YOY Giant Sea Bass show through the sand to the left of
the insert.

the exact spot where the GSB had disappeared (Fig. 3), and then put a 15 cm net over the
spot, working the net’s frame down into the sand. The fish remained buried. I dug my
hand deep into the soft sand under the net, lifted a ball of sand containing the fish up into
the net. As the sand fell away to the sides of the ball, the GSB burst out of the sand and
up into the net. I measured the GSB at 32 mm, and released it. The GSB darted beneath
me as I knelt on the sand. I pushed off the bottom, but the fish was gone.

The GSB’s actions appeared to be a flight response, a possible last-ditch effort to evade
predation in an area where there was no available cover for shelter. Unlike flatfish that may
cover themselves with soft sediment to ambush their prey (Gibson and Robb 1991 ), or certain
benthic gobiid fishes that have mutualistic relationships with shrimp that dig holes for
shelters (Horinouchi 2008; Thacker et. al. 2011), this GSB was behaving as if actively trying
to avoid detection by a “predator”. Senoritas, Oxyjulis calif ornica, are also known to bury
themselves to avoid predators, but this occurs mostly at night, with the senorita remaining
buried for protection (Hobson, E.S. 1968), and not as an immediate escape response. This
GSB predator evasion method should prove highly effective, as few predators could dig in the
sand for the fish after burial. The bottom was so uniform that if the observer had looked
away from the spot where the fish had buried, its location would have been lost (Fig. 3).

The open expanse of sand and mud bottoms that these small YOY GSB utilize make
ideal nursery areas since there is an abundance of food, such as mysids and other
arthropods (Dahl 1952), and relatively few inhabitants, including predators (McLachlan
1990). However, when a juvenile does encounter a predator, it must rely on its cryptic
coloration and shape, and other types of active and passive predator avoidance strategies.
The burying behavior observed may be a successful last-resort predator avoidance
strategy for GSB, and is certainly the first one documented.

A POSSIBLE PREDATOR AVOIDANCE MECHANISM

Acknowledgements

We would like to thank R. H. Defran of San Diego State University, L. G. Allen of
California State University, Northridge, and D. J. Pondella II for reviewing this note.

Literature Cited

Baldwin, D.S. and Kaiser, A. 2008. Giant sea bass, Stereolepis gigas, status of the fisheries report. Cal.
Dept. Fish Game. p. 8.

Dahl, E. 1952. Some aspects of the ecology and zonation of the fauna on sandy beaches. Oikos, 4(l):l-27.
Hobson, E.S. 1965. Diurnal-Nocturnal Activity of Some Inshore Fishes in the Gulf of California. Copeia,
291-302.

Horinouchi, M. 2008. Patterns of food and microhabitat resource use by two benthic Gobiid fishes.
Environ. Biol. Fish., 82:187-194.

Gibson, R.N. and Robb, D.L. 1992. The relationship between body size, sediment grain size and the
burying ability of juvenile plaice, Pleuronectes platessa. L. J. Fish Biol., 40(77):1— 778.

McLachlan, A. 1990. Dissipative beaches and macrofauna communities on exposed intertidal sands.
J. Coast. Res., 6(1):57— 71.

Miller, D.J. and Lea, R.N. 1972. Guide to the Coastal Marine Fishes of California. Calif. Dept. Fish.
Game, Fish Bull., 157-249.

Pondella, D.J. II and Allen, L.G. 2008. The decline and recovery of four predatory fishes from the
Southern California Bight. Mar. Biol., 154:307-313.

Thacker, C., Thompson, A., and Roje, D. 2011. Phylogeny and evolution of Indo-Pacific shrimp-
associated gobies Gobiiformes: Gobiidae. Mol. Phylog. Evol., 59( 1):168— 176.

Bull. Southern California Acad. Sci.

114(1), 2015, pp. 58-62

Nelson’s big horn sheep ( Ovis canadensis nelsoni ) trample
Agassiz’s desert tortoise ( Gopherus agassizii ) burrow at
a California wind energy facility

Mickey Agha,1 David Delaney,2 Jeffrey E. Lovich,3 Jessica Briggs,4 Meaghan Austin3

and Steven J. Price1

1 Department of Forestry, University of Kentucky, Lexington, KY 40546, USA
2U.S. Army Construction Engineering Research Laboratory, P.O. Box 9005,
Champaign, IL 61826-9005, USA

3 US. Geological Survey, Southwest Biological Science Center, 2255 North Gemini
Drive, MS-9394, Flagstaff, Arizona 86001, USA
4 Colorado State University, Fort Collins, CO 80523, USA

Research on interactions between Agassiz’s desert tortoises ( Gopherus agassizii ) and
ungulates has focused exclusively on the effects of livestock grazing on tortoises and their
habitat (Oldemeyer, 1994). For example, during a 1980 study in San Bernardino County,
California, 164 desert tortoise burrows were assessed for vulnerability to trampling by
domestic sheep ( Ovis aries). Herds of grazing sheep damaged 10% and destroyed 4% of
the burrows (Nicholson and Humphreys 1981). In addition, a juvenile desert tortoise was
trapped and an adult male was blocked from entering a burrow due to trampling by
domestic sheep. Another study found that domestic cattle ( Bos taurus) trampled active
desert tortoise burrows and vegetation surrounding burrows (Avery and Neibergs 1997).
Trampling also has negative impacts on diversity of vegetation and intershrub soil crusts
in the desert southwest (Webb and Stielstra 1979). Trampling of important food plants
and overgrazing has the potential to create competition between desert tortoises and
domestic livestock (Berry 1978; Coombs 1979; Webb and Stielstra 1979).

Native ungulates like Nelson’s big horn sheep ( Ovis canadensis nelsoni ) co-occur with
desert tortoises in portions of the desert southwest. Due to habitat and partial dietary
overlap of various annual forbs and grasses at certain elevations (Ernst and Lovich 2009;
Oehler et al. 2003), there is potential for contact between these species. Although there are
data demonstrating damage and destruction of desert tortoise burrows caused by
domestic ungulates (Nicholson and Humphreys 1981; Avery and Neibergs 1997), it is
previously undocumented if native sheep like Nelson’s big horn sheep are capable of
similar impacts to tortoise burrows.

On 29 September 2013, we documented desert tortoise burrow collapse caused by
Nelson’s big horn sheep trampling at a wind energy facility in Riverside Co., California,
USA (33°57'06"N, 116°40,02,'W, WGS84). In the summer of 2013 (1 June 2013 to 14
November) 48 Reconyx and Wildgame motion sensor trail cameras were deployed at the
entrances of desert tortoise burrows during an ongoing investigation of the effects of
wind energy generation on behavior and activity of this species (Lovich et al. 2014).
Cameras were mounted on 1.5 m foot tall steel stakes inserted into the ground
approximately 1 m from desert tortoise burrow entrances that were known to be occupied
or used recently. Cameras were activated by movement of wildlife via an infrared sensor,

Corresponding author: steven.price@uky.edu

BIG HORN SHEEP AND TORTOISES

Fig. 1 . Active desert tortoise burrow collapse caused by Nelson’s big horn sheep in a series of 4 motion
sensor camera images.

and programmed to take 1-5 photographs at a trigger speed of 0.2 sec. Each month, an
investigator checked each camera and downloaded photos onto a data storage device.
Lastly, surface air temperatures were collected every 30 minutes from an onsite Remote
Automated Weather Station (WWAC1; accessed via the MesoWest website (http://
mesowest.utah.edu/index.html).

Our motion sensor cameras recorded three Nelson’s big horn sheep approach a north
facing active desert tortoise burrow (previously occupied by an adult male desert tortoise
on 13 June 2013) at 1140 h. From 1241 h to 1303 h, several Nelson’s big horn sheep
gathered below the entrance of the burrow, brushing loosened soil around the entrance of
the burrow with their hooves, eventually causing the outer walls to collapse (Fig 1 .). Three
different Nelson’s big horn sheep then proceeded to lie down and in the process compact
the soil, rocks and sticks on top of the newly collapsed entrance from 1304 h to 1429 h.

Fig. 2. Nelson’s big horn sheep at the entrance of desert tortoise burrow. Nelson’s big horn sheep may
have been eating tortoise feces or soil, or simply investigating the burrow.

SOUTHERN CALIFORNIA ACADEMY OF SCIENCES

2013-09-02 11:26:01

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Fig. 3. Domestic cattle walking past the entrance of a desert tortoise burrow.

Several Nelson’s big horn sheep remained standing at the burrow from 1430 h to 1542 h.
During these observations ambient air temperature ranged from 30.56 C to 32.50 C.

Over the course of the camera-trapping study, Nelson’s big horn sheep were also
recorded walking past or standing at the base of seven different desert tortoise burrows at
various other locations throughout the study site. Photographs also revealed what
appeared to be Nelson’s big horn sheep grazing near the mouth of the burrow (Fig. 2).
Since few plants grow in the mouth of active tortoise burrows, the sheep may have been
eating soil or possibly the fresh feces of desert tortoises that are comprised mostly of
partially digested grass and forbs. In addition to big horn sheep, domestic cattle were
captured by motion sensor cameras walking past the base of four desert tortoise burrows
(Fig. 3).

The images we recorded are the first documented evidence of Nelson’s big horn sheep
trampling a desert tortoise burrow and subsequently collapsing the outer walls of the
burrow in the process. Nelson’s big horn sheep employ various strategies of seeking shade
and cooler soil for bedding (Cain et al. 2008), and it appears that north-facing slopes
(location of collapsed tortoise-burrow) may provide such a site. Alternatively, previous
studies of big horn sheep have documented extensive movement and occasionally large
descents from mountain ranges to use mineral licks at lower elevations, as they provide
sodium which is crucial to physiological functions (Bangs et al. 2005; Holl and Bleich
1987; Watts and Schemnitz 1985). Since most terrestrial plants have low concentrations
of sodium (Weeks and Kirkpatrick 1976), Nelson’s big horn sheep may be mining
essential minerals brought to the surface by tortoises through excavation of their burrows
(Ernst and Lovich 2009; Turner et al. 1984). One study demonstrating soil ingestion, or
geophagy, by bighorn sheep ( Ovis canadensis) in Alberta, Canada found their feces
contained as much as 30% soil in some samples (Skipworth 1974). Ingestion of desert
tortoise burrow soil may be important to Nelson’s big horn sheep as it could be a source
of certain minerals (Beyer et al. 1994). Lastly, we hypothesize that relatively high plant
productivity at the site (Ennen et al. 2012b; Lovich et al. 2015) attracts ungulates (Oehler
et al. 2003), both domestic and native to the study area. Moderate winter precipitation
produces an abundance of annual food plants for both desert tortoises and big horn
sheep at the study site.

Trampling and collapsing active desert tortoise burrows may entomb resident
individuals (Loughran et al. 2011; Nichols and Humphries 1981), although in the
majority of observed burrow collapses at the site, tortoises were able to excavate

BIG HORN SHEEP AND TORTOISES

themselves (Loughran et al. 2011). In light of our observation, trampling may have
greater impacts to slope dwelling rather than valley dwelling desert tortoises.
Furthermore, female desert tortoises nest at the entrance and within burrows (Agha et
al. 2013; Ennen et al. 2012a); consequently, trampling may negatively impact tortoise egg
clutches or entomb emerging neonates (Berry 1978). Entombment of desert tortoises
within burrows can cause physiological stress to the animal (Loughran et al. 2011),
thereby leading to potential mortality (Lovich et al. 2011). We are unaware of any cases
where bighorn sheep behavior resulted in mortality of desert tortoises and suspect that
such interactions between the species are rare in comparison to interactions involving
domestic ungulates.

Acknowledgements

Our research was supported by the California Energy Commission-Public Interest
Energy Research Program (Contract NO.: 500-09-020), the California Desert District
Office of the Bureau of Land Management, U.S. Army Construction Engineering
Research Laboratory, and the Desert Legacy Fund of the California Desert Research
Program. Research was conducted under permits from the United States Fish and
Wildlife Service, California Department of Fish and Game, and the Bureau of Land
Management. Earlier versions of the manuscript benefited greatly from comments
offered by Vernon Bleich. Special thanks are given to A. Muth of the Boyd Deep Canyon
Desert Research Center of the University of California, Riverside, for providing
accommodations during our research. Any use of trade, product, or firm names is for
descriptive purposes only and does not imply endorsement by the U.S. Government.

Literature Cited

Agha, M., J.E. Lovich, J.R. Ennen, and E. Wilcox. 2013. Nest-guarding by female Agassiz’s desert tortoise
{Gopherus agassizii ) at a wind-energy facility near Palm Springs, California. The Southwestern
Naturalist, 58:254-257.

Avery, H.W., and A.G. Neibergs. 1997. Effects of cattle grazing on the desert tortoise, Gopherus agassizii :
nutritional and behavioral interactions. In Proceedings: Conservation, Restoration, and Manage-
ment of Tortoises and Turtles- An International Conference, 13-20.

Bangs, P.D., P.R. Krausman, K.E. Kunkel, and Z.D. Parsons. 2005. Habitat use by female desert bighorn
sheep in the Fra Cristobal Mountains, New Mexico, USA. European Journal of Wildlife Research,
51:77-83.

Berry, K.H. 1978. Livestock grazing and the desert tortoise. In Transactions of the North American
Wildlife and Natural Resources Conference (USA), 1978:505-519.

Beyer, W.N., E.E. Connor, and S. Gerould. 1994. Estimates of soil ingestion by wildlife. Journal of
Wildlife Management, 58:375-382.

Cain, J.W., B.D. Jansen, R.R. Wilson, and P.R. Krausman. 2008. Potential thermoregulatory advantages
of shade use by desert bighorn sheep. Journal of Arid environments, 72:1518-1525.

Coombs, E.M. 1979. Food habits and livestock competition with the desert tortoise on the Beaver Dam
Slope, Utah. Proceedings of the Desert Tortoise Council, 1979:132-147.

Ennen, J.R., J.E. Lovich, K.P. Meyer, C. Bjurlin, and T.R. Arundel. 2012a. Nesting Ecology of
a Population of Gopherus agassizii at a Utility-Scale Wind Energy Facility in Southern California.
Copeia, 2012:222-228.

Ennen, J.R., K. Meyer, and J.E. Lovich. 2012b. Female Agassiz’s desert tortoise activity at a wind energy
facility in southern California: The influence of an El Nino event. Natural Science, 4:30-37.
Ernst, C.H., and Lovich, J.E. 2009. Turtles of the United States and Canada. Johns Hopkins University
Press.

Holl, S.A., and V.C. Bleich. 1987. Mineral lick use by mountain sheep in the San Gabriel Mountains,
California. Journal of Wildlife Management, 51:383-385.

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Lovich, J.E., D. Delaney, J. Briggs, M. Agha, M. Austin, and J. Reese. 2014. Black bears ( Ursus
americanus ) as a novel potential predator of Agassiz’s desert tortoises ( Gopherus agassizii ) at
a California wind energy facility. Bulletin of the Southern California Academy of Sciences, 113:
34-41.

, J.R. Ennen, K. Meyer, M. Agha, C. Loughran, C. Bjurlin, M. Austin, S. Madrak. 2015. Not

putting all their eggs in one basket: bet-hedging despite extraordinary annual reproductive output
of desert tortoises. Biological Journal of the Linnean Society, 115.2:399^-10.

, J.R. Ennen, S.V. Madrak, and B. Grover. 2011. Turtles, culverts and alternative energy

development: an unreported but potentially significant mortality threat to the desert tortoise
( Gopherus agassizii ). Chelonian Conservation and Biology, 10:124-129.

Loughran, C.L., J.E. Ennen, and J.E. Lovich. 201 1 . Gopherus agassizii (Desert tortoise). Burrow collapse.
Herpetological Review, 42:593.

Nicholson, L., and K. Humphreys. 1981. Sheep grazing at the Kramer study plot, San Bernardino County,
California. In Proceedings of the 1981 symposium of the Desert Tortoise Council, 163-194.

Oehler Sr., M.W., R.T. Bowyer, and Y.C. Bleich. 2003. Home ranges of female mountain sheep, Ovis
canadensis nelsoni: effects of precipitation in a desert ecosystem. Mammalia, 67:385-402.

Oldemeyer, J.L. 1994. Livestock grazing and the desert tortoise in the Mojave Desert, p. 95-103. In R. B.
Bury and D. J. Germano (eds.), Biology of North American tortoises. U.S. Dept. Int. Natl. Biol.
Surv. Fish Wildl. Res. 13.

Skipworth, J.P. 1974. Ingestion of grit by bighorn sheep. Journal of Wildlife Management, 38:880-883.

Turner, F.B., P. A Medica, and C.L. Lyons. 1984. Reproduction and survival of the desert tortoise
(Scaptochelys agassizii) in Ivanpah Valley, California. Copeia, 1984:811-820.

Watts, T.J., and S.D. Schemnitz. 1985. Mineral lick use and movement in a remnant desert bighorn sheep
population. Journal of Wildlife Management, 49:994-996.

Webb, R.H., and Stielstra, S.S. 1979. Sheep grazing effects on Mojave Desert vegetation and soils.
Environmental Management, 3:517-529.

Weeks Jr., H.P., and C.M. Kirkpatrick, 1976. Adaptations of white-tailed deer to naturally occurring
sodium deficiencies. Journal of Wildlife Management, 40:610-625.

SMITHSONIAN LIBRARIES

3 9088 01817 2296

CONTENTS

Possible Stock Structure of Coastal Bottlenose Dolphins off Baja California
and California Revealed by Photo-Identification Research. R.H. Defran,
Marthajane Caldwell, Eduardo Morteo, Aimee R. Lang, Megan G. Rice, and
David W. Weller ... J 1

Removal Efforts and Ecosystem Effects of Invasive Red Swamp Crayfish
(Procambarus clarkii) in Topanga Creek, California. Crystal Garcia, Elizabeth
Montgomery, Jenna Krug, and Rosi Dagit 12

Soil Organic Carbon and Nitrogen Storage in Two Southern California Salt Marshes:

The Role of Pre-Restoration Vegetation. Jason K. Keller, Tyler Anthony,
Dustin Clark, Kristin Gabriel, Dewmini Gamalath, Ryan Kabala, Julie King,
Ladyssara Medina, and Monica Nguyen 22

Identical Response of Caged Rock Crabs (Genera Metacarcinus and Cancer )
to Energized and Unenergized Undersea Power Cables in Southern
California, USA. Milton S. Love, Mary M. Nishimoto, Scott Clark, and
Ann Scarborough Bull 33

Recent Decline of Lowland Populations of the Western Gray Squirrel in the
Los Angeles Area of Southern California. Daniel S. Cooper and Alan E.
Muchlinkski 42

A Young-of-the-Year Giant Sea Bass, Stereolepis gigas Buries Itself in Sandy
Bottom: A Possible Predator Avoidance Mechanism. Michael C. Couffer and
Stephanie A. Benseman 54

Nelson’s Big Horn Sheep (Ovis Canadensis nelsoni ) Trample Agassiz’s Desert
Tortoise ( Gopherus agassizii) Burrow at a California Wind Energy Facility.
Mickey Agha, David Delaney, Jeffrey E. Lovich, Jessica Briggs, Meaghan
Austin, and Steven J. Price 58

Cover: Young-of-the-Year Giant Sea Bass from Newport Beach, California. Photo by permission of
Michael C. Couffer.

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